Optimization of sustainable nitrification strategies for ...lib.ugent.be/fulltxt/RUG01/002/166/674/RUG01-002166674_2014_000… · enorm leerrijke ervaring en een pracht van een afsluiter

Faculteit Bio-ingenieurswetenschappen

Academiejaar 2013 - 2014

Optimization of sustainable nitrification strategies

for source separated urine

Maarten Muys

Promotor: Prof. dr. ir. Nico Boon

Co-promotor: Prof. dr. ir. Siegfried E. Vlaeminck

Tutor: Ing. Joeri Coppens

Scriptie voorgedragen tot het behalen van de graad van

MASTER IN DE BIO-INGENIEURSWETENSCHAPPEN: MILIEUTECHNOLOGIE

Copyright

“De auteur en de promotor geven de toelating deze scriptie voor consultatie beschikbaar te stellen

en delen ervan te kopiëren voor persoonlijk gebruik. Elk ander gebruik valt onder de beperkingen van

het auteursrecht, in het bijzonder met betrekking tot de verplichting de bron te vermelden bij het

aanhalen van resultaten uit deze scriptie.”

“The author and the promoter give the permission to use this thesis for consultation and to copy

parts of it for personal use. Every other use is subject to the copyright laws, more specifically the

source must be extensively specified when using results from this thesis.”

Gent, 5 juni 2014

De promotor, De co-promotor, De auteur,

Prof. dr. ir. Nico Boon Prof. dr. ir. Siegfried E. Vlaeminck Maarten Muys

Woord vooraf

i

WOORD VOORAF

“The most beautiful thing we can experience is the mysterious. It is the source of all true art and all

science.”

Albert Einstein

Beste lezer,

Het afgelopen jaar was er één van uitdagingen. Een jaar onderzoek verrichten aan LabMET was een

enorm leerrijke ervaring en een pracht van een afsluiter na 5 jaren studie aan dat geliefde

“Boerekot”. Deze masterproef had ik echter nooit kunnen realiseren zonder de hulp, raad en

aanmoediging van velen die ik hier graag wil bedanken.

In de eerste plaats wil ik mijn twee enthousiaste promotoren Prof. dr. ir. Nico Boon en Prof. dr. ir.

Siegfried E. Vlaeminck bedanken voor hun wijze raad waarmee ze klaarstonden en de kritische kijk en

ideeën die ze gaven tijdens de maandelijkse meetings.

Naast mijn twee promoteren wil ik ook mijn tutor Joeri Coppens bedanken die, ondanks zijn drukke

agenda, toch dagelijks klaar stond om uitleg te geven en bij te sturen waar nodig. Zijn ruime kennis

over alles wat stikstof betreft en kritische kijk naar onderzoek toe hebben mij geïnspireerd om dit

werk tot een goed einde te brengen. Naast Joeri bedank ik ook Peter, Sam, Francis, Alberto, Marta en

Emilie voor het advies in het labo.

Natuurlijk wordt het dagelijkse leven bij LabMET in goede banen geleid door een heleboel mensen

waaronder ik Mike, Greet en Robin in het bijzonder wil bedanken voor hun hulp.

Heel de groep thesisstudenten, waaronder ik ondertussen vele vrienden tel, wil ik bedanken voor

deze mooie tijd. Wanneer iets fout liep was er altijd wel iemand die klaar stond met steun en raad. In

het bijzonder dank ik Chaïm, Maarten, Ruben, Tim, Lien, Charles, Wim, Stijn, Thomas, Kristof, Emmy

en Daniel voor de leuke momenten in het labo en erbuiten!

Mijn familie, die onvoorwaardelijk en altijd voor mij klaar staat geniet mijn grootste dank. Zonder

hen was studeren nooit een mogelijkheid geweest. Tot slot en niet in het minst bedank ik mijn lieve

vriendin, Jolien, voor haar liefde, steun en luisterend oor.

Maarten Muys – Juni 2014

Table of Contents

ii

TABLE OF CONTENTS

Woord vooraf ....................................................................................................................................... i

Table of Contents ................................................................................................................................ ii

Used abbreviations ..............................................................................................................................vi

List of Figures ...................................................................................................................................... vii

List of Tables ........................................................................................................................................ ix

Abstract ............................................................................................................................................... x

PART 1: Literature study ......................................................................................................................... 1

1. Introduction ..................................................................................................................................... 1

1.1. Why nutrient recovery? ........................................................................................................... 1

1.2. Urine ........................................................................................................................................ 3

1.2.1. Composition and properties ................................................................................................ 3

1.2.2. Opportunities for nutrient recovery .................................................................................... 4

2. Nitrification ...................................................................................................................................... 7

2.1. General principles .................................................................................................................... 7

2.2. Process conditions for nitrification .......................................................................................... 8

2.2.1. Free ammonia and free nitrous acid inhibition ............................................................... 8

2.2.2. pH..................................................................................................................................... 9

2.2.3. Oxygen ........................................................................................................................... 10

2.2.4. Temperature .................................................................................................................. 10

2.2.5. Salinity ........................................................................................................................... 10

2.2.6. Light ............................................................................................................................... 11

2.3. Nitrification of urine .............................................................................................................. 11

2.4. Membrane Bioreactor ........................................................................................................... 13

2.5. Aerobic sludge granulation .................................................................................................... 14

3. Microalgae and their applications in wastewater treatment........................................................ 15

3.1. Microalgae ............................................................................................................................. 15

3.2. Application of microalgae in a WWTP ................................................................................... 15

3.3. Photosynthetic aeration ........................................................................................................ 17

3.4. Algae on urine ........................................................................................................................ 17

Table of contents

iii

4. Research objectives ....................................................................................................................... 18

PART 2: Materials and Methods .......................................................................................................... 19

1. Photobioreactor bubble column ................................................................................................... 19

1.1. Reactor design and set up ..................................................................................................... 19

1.2. Operation .............................................................................................................................. 21

1.3. Synthetic urine ...................................................................................................................... 22

1.4. Sludge and algae ................................................................................................................... 23

2. Membrane bioreactors ................................................................................................................. 25

2.1. Reactor set up ....................................................................................................................... 25

2.2. Inoculum ................................................................................................................................ 26

2.3. Operation .............................................................................................................................. 26

2.4. Synthetic hydrolyzed urine.................................................................................................... 27

3. Batch experiments ........................................................................................................................ 27

3.1. Effect of salt concentration on nitrification activity ............................................................. 27

3.2. Effect of ammonium and salt concentration on algae growth ............................................. 28

3.3. Nitrification activity test protocol ......................................................................................... 28

4. Analytical techniques .................................................................................................................... 29

4.1. Ammonium ............................................................................................................................ 29

4.1.1. Teststrip ......................................................................................................................... 29

4.1.2. Nessler colorimetric method ......................................................................................... 29

4.1.3. Steam distillation ........................................................................................................... 29

4.2. Total Kjeldahl nitrogen .......................................................................................................... 30

4.3. Nitrite and Nitrate ................................................................................................................. 30

4.3.1. Teststrip ......................................................................................................................... 30

4.3.2. Ion chromatography ...................................................................................................... 30

4.4. Chemical oxygen demand ..................................................................................................... 30

4.5. pH .......................................................................................................................................... 30

4.6. Dissolved oxygen ................................................................................................................... 31

4.7. Electrical conductivity ........................................................................................................... 31

4.8. Total suspended solids and Volatile suspended solids ......................................................... 31

4.9. Headspace: Oxygen, nitrogen and carbon dioxide ............................................................... 31

5. Microscopy .................................................................................................................................... 31

Table of contents

iv

PART 3: Results ..................................................................................................................................... 32

1. Photobioreactor (PBR) ................................................................................................................... 32

1.1. Preliminary batch tests .......................................................................................................... 32

1.1.1. Effect of salt concentration on nitrification activity ...................................................... 32

1.1.2. Effect of ammonium and salt concentration on algae growth ...................................... 33

1.2. PBR Nitrification reactor ........................................................................................................ 34

1.2.1. Dissolved oxygen profile over one cycle ....................................................................... 41

1.3. Microscopy on PBR biomass .................................................................................................. 41

2. Membrane bioreactors (MBR) ....................................................................................................... 43

2.1. Nitrification activity test B-Sludge ......................................................................................... 43

2.2. MBR Nitrification reactors ..................................................................................................... 44

2.2.1. MBR 1: High salinity start-up strategy ........................................................................... 44

1.1.1. MBR 2: Increasing salinity start-up strategy .................................................................. 49

PART 4: Discussion ................................................................................................................................ 55

1. Photobioreactor ............................................................................................................................ 55

1.1. Nitrification efficiency ........................................................................................................... 55

1.2. Reactor operation parameters .............................................................................................. 57

1.3. Photo-aeration ...................................................................................................................... 58

1.4. Reactor short-comings .......................................................................................................... 59

1.4.1. Algae and AOB/NOB growth .......................................................................................... 59

1.4.2. Settling ........................................................................................................................... 62

1.4.3. Photo-inhibition ............................................................................................................. 62

1.5. Reactor improvement possibilities ........................................................................................ 63

1.6. Microscopy ............................................................................................................................ 65

2. Membrane bioreactors .................................................................................................................. 65

2.1. Influence of the medium ....................................................................................................... 65

2.2. MBR 1: High salinity start-up strategy ................................................................................... 66

2.3. MBR 2: Increasing salinity start-up strategy .......................................................................... 67

2.4. Comparison with literature ................................................................................................... 68

2.5. Adaptation to higher salinities .............................................................................................. 68

2.6. Microscopy ............................................................................................................................ 69

3. Influence of salt concentration on nitrification activity ................................................................ 69

Table of contents

v

4. Influence of salinity and ammonium concentration on the microalgal growth ........................... 72

5. Economical point of view .............................................................................................................. 73

6. Conclusions ................................................................................................................................... 73

7. Further research ............................................................................................................................ 75

7.1. Photo-aeration for urine stabilization ................................................................................... 75

7.2. Urine stabilization with an MBR configuration ..................................................................... 75

References ............................................................................................................................................. 76

Appendix ............................................................................................................................................... 86

1. Effect of ammonium concentration on different microalgal species ....................................... 86

2. Effect of salt concentration on different microalgal species .................................................... 87

Used abbreviation

vi

USED ABBREVIATIONS

AOB Ammonium Oxidizing Bacteria

COD Chemical Oxygen Demand

CSTR Continuously Stirred Tank Reactor

DO Dissolved Oxygen

EC Electrical Conductivity

FA Free Ammonia

FNA Free Nitrous Acid

HRAP High Rate Algal Pond

HRT Hydraulic Retention Time

IC Ion Chromatography

MBR Membrane Bioreactor

NOB Nitrite Oxidizing Bacteria

OD Optical Density

PBR Photobioreactor

SBR Sequential Batch Reactor

SRT Sludge Retention Time

TKN Total Kjeldahl Nitrogen

TSS Total Suspended Solids

VSS Volatile Suspended Solids

WWTP Wastewater Treatment Plant

List of Figures

vii

LIST OF FIGURES

Figure 1. The dependence of the NH4+/NH3 ratio as a function of pH (Clegg and Whitfield, 1995). ...... 9

Figure 2. Left: External loop process. Right: Submerged membrane process (Carstensen et al., 2012).

............................................................................................................................................................... 14

Figure 3. Reactor Setup: schematic overview of the lab-scale photobioreactor bubble column. ........ 20

Figure 4. Reactor setup: lab-scale photobioreactor bubble column. ................................................... 20

Figure 5. Timeline for 1 operation cycle of the PBR. ............................................................................. 21

Figure 6. Algae cultivation before reactor inoculation. ........................................................................ 23

Figure 7. Reactor Setup: Schematic overview of the lab-scale membrane bioreactor......................... 25

Figure 8. Reactor Setup: Lab-scale membrane bioreactors. ................................................................. 26

Figure 9. Division of the 96 well micro titer plates. .............................................................................. 28

Figure 10. Effect of different salt concentrations on HANDS activity. Top left: 1 g NaCl L-1. Top right: 2

g NaCl L-1. Bottom left: 3.5 g NaCl L-1. Bottom right: 5 g NaCl L-1. ......................................................... 33

Figure 11. OD at 620 nm in function of time for different salt and ammonium concentrations Left:

effect of salinity on the algal mix; Right: effect of ammonium concentration on the algal mix. ......... 34

Figure 12. Nitrogen species in influent and effluent, nitrogen loading rate (gN L-1 d-1) and nitrification

efficiency (%) during PBR operation. ..................................................................................................... 35

Figure 13. Influent and effluent COD concentrations and COD removal efficiency during PBR

operation. .............................................................................................................................................. 36

Figure 14. Dissolved oxygen, TSS and VSS concentration in function of time during PBR operation. .. 36

Figure 15. Ammonium, nitrite and nitrate profiles. Upper left: HANDS + reactor supernatans; Upper

right: PBR biomass + fresh medium; Bottom: PBR biomass + reactor supernatans. ............................ 37

Figure 16. Activity test on the PBR biomass after 146 days. ................................................................. 39

Figure 17. Activity test on the PBR biomass after 180 days. ................................................................. 40

Figure 18. Dissolved oxygen profile during 1 cycle. .............................................................................. 41

Figure 19. Microscopy images of PBR biomass. Top: 400 x magnification; Bottom: 1000 x

magnification. (1) Chlorella sp.; (2) Scenedesmus sp.; (3) Synechocystis sp.; (4) Leptolyngbya sp.; (5)

Bacteria. ................................................................................................................................................ 42

Figure 20. Nitrogen species profiles of the preliminary activity test on inoculum B-sludge. ............... 43

Figure 21. Nitrogen species in the effluent for MBR 1. ......................................................................... 44

Figure 22. Reactor free ammonia (FA) concentrations for MBR 1. ....................................................... 45

Figure 23. Nitrogen loading rate, influent urine dilution and EC in reactor effluent for MBR 1. ......... 45

Figure 24. Left: TSS and VSS concentrations and the VSS/TSS ratio. Right: Nitrification efficiency. .... 46

List of Figures

viii

Figure 25. Results nitrification activity test. Left: MBR I, 5 mS cm-1; Right: MBR I, 70 mS cm-1. ........... 46

Figure 26. Results nitrification activity test. Top left: MBR I, 5 mS cm-1; Top right: MBR I, 15 mS cm-1;

Bottom: MBR I, 45 mS cm-1; .................................................................................................................. 48

Figure 27. AOB and NOB activity trend at different salinities for MBR 1. ............................................. 49

Figure 28. Microscope images of biomass MBR 1. Left: 400 x magnification; Right: 1000 x

magnification. ........................................................................................................................................ 49

Figure 29. Nitrogen species in the effluent for MBR 2. ......................................................................... 50

Figure 30. Reactor free ammonia (FA) concentrations for MBR 2. ....................................................... 50

Figure 31. Nitrogen loading rate, influent urine dilution and electric conductivity for MBR 2. ........... 51

Figure 32. Left: TSS and VSS concentrations and the VSS/TSS ratio. Right: Nitrification efficiencies. .. 51

Figure 33. Results nitrification activity test. Left: MBR II, 5 mS cm-1; Right: MBR II, 27 mS cm-1. ......... 52

Figure 34. Results nitrification activity test. Top left: MBR II, 5 mS cm-1; Top right: MBR II, 15 mS cm-1;

Bottom: MBR II, 45 mS cm-1. ................................................................................................................. 53

Figure 35. AOB and NOB activity trend at different salinities for MBR 2. ............................................. 54


magnification. ........................................................................................................................................ 54

Figure 37. Theoretical equilibrium between microalgae and nitrifiers starting from 100 units nitrogen

and 90 units COD (average ratio in urine). ............................................................................................ 59

Figure 38. Light distribution inside a photobioreactor containing 1 g L−1 Euglena gracilis cells with a

light absorption coefficient of 200 m2 kg−1. The photobioreactor was illuminated from one surface at

an intensity of 500 µmol m-2 s−1. ........................................................................................................... 60

Figure 39. Settling phase in the reactor after reïnoculation with ABIL and HANDS. ............................ 61

Figure 40. Foam formation MBR 1. ....................................................................................................... 66

Figure 41. Profile of specific removal of organic (DOC) and NH4+-N at different salt concentrations. 70

Figure 42. Growth curves of different microalgae species at different ammonium concentrations. A.

Chlorella sp.; B. Haematococcus sp.; C. Desmodesmus sp.; D. Ankistrodesmus sp.; E. Pediastrum

duplex; F. Chlorella vulgaris; G. Algae mix from grassland pond; H. Nannochloropsis sp. .................... 87

Figure 43. Growth curves of different microalgae species at different salt concentrations. A. Chlorella

sp.; B. Haematococcus sp.; C. Desmodesmus sp.; D. Ankistrodesmus sp.; E. Pediastrum duplex; F.

Chlorella vulgaris; G. Algae mix from grassland pond; H. Nannochloropsis sp. .................................... 88

List of Tables

ix

LIST OF TABLES

Table 1. Agriculture production and resource use in the recent past to the near future (Vance, 2001).

................................................................................................................................................................. 2

Table 2. Composition of urine (Udert et al., 2003b). .............................................................................. 4

Table 3. Inhibition of AOB and NOB by free ammonia (FA) and free nitrous acid (FNA) (Anthonisen et

al., 1976). ................................................................................................................................................. 9

Table 4. Comparison of reactor tests for urine nitrification. ................................................................ 12

Table 5. Comparison of open and closed systems for microalgae systems +: advantage; -:

disadvantage (Van Den Hende, 2014). .................................................................................................. 16

Table 6. Overview of the initial and final photobioreactor operation parameters. ............................. 22

Table 7. Composition of synthetic urine. .............................................................................................. 22

Table 8. ESAW culture medium (Harrison et al., 1980)......................................................................... 24

Table 9. Overview of the initial operation parameters (Start-up*/Target values**). .......................... 27

Table 10. Composition of synthetic hydrolyzed urine. ......................................................................... 27

Table 11. Nitrification activities of HANDS sludge for different electrical conductivities. .................... 32

Table 12. Nitrification activities for the three different configurations................................................ 38

Table 13. PBR biomass nitrification activities after 146 days and after 180 days, pH during the

experiment and initial VSS concentration. ............................................................................................ 40

Table 14. Nitrification activity B-sludge, pH and DO during the experiment and initial VSS

concentration. ....................................................................................................................................... 43

Table 15. Nitrification activity MBR 1 different conductivities at day 21. ............................................ 47




Table 19. Typical values of substrate affinity constants (Ks) and doubling time for AerAOB, NOB, and

Microalgae species (Guisasola et al., 2007; Hein et al., 1995; Lackner et al., 2008). ‘-‘: not applicable.

............................................................................................................................................................... 61

Table 20. Nitrosomonas respiration evaluation (Alleman et al. 1987). ................................................ 63

Table 21. Nitrification activities at different salinities. ......................................................................... 71

Table 22. Nitrification activity MBR 2 at different temperatures (after 44 days). ................................ 72

Abstract

x

ABSTRACT

Human urine accounts for only 1% of the total wastewater volume, but it contains approximately

40% of the phosphorus load, 69% of the nitrogen load, and 60% of the potassium load that arrives in

a municipal wastewater treatment plant. A more sustainable alternative to replace the conventional

and inefficient biological nitrogen and chemical phosphorous removal might lie in the recovery of

these valuable nutrients. Source separated urine however, is a very unstable waste stream in non-

sterile conditions. During collection and transportation, bacterial urease hydrolyses the organic

nitrogen present which results in a pH rise and subsequent ammonia volatilization. This volatilization

results in unwanted nitrogen loss and odor problems. Biological oxidation of the present nitrogen to

nitrate is a possible solution to tackle this problem and stabilize urine.

A first part of this thesis deals with the aeration cost for nitrification. Mechanical aeration accounts

for 45 – 75% of the total energy consumption in biological wastewater treatment plants. An

alternative for mechanical aeration is the in-situ production of oxygen by photosynthetic organisms,

or photo-aeration. A first goal of this research project is to develop a consortium of microalgae and

nitrifying bacteria, in which microalgae produce the oxygen in-situ which is used by the nitrifiers to

oxidize ammonium. This process is studied in a bubble column photobioreactor. Nitrification initiated

after 55 days of operation and the maximum reached nitrification efficiency was 69% (total influent

nitrogen that was converted to nitrate) at a nitrogen loading rate of 90 mg N L-1 d-1 (12% dilution of

synthetic urine). COD removal efficiencies obtained were situated between 44 and 83%.

The high nitrogen concentration, salinity and free ammonia in 100% hydrolyzed urine limit the

activity of non-adapted ammonium oxidizing bacteria (AOB) and nitrite oxidizing bacteria (NOB). A

second goal of this research project is to assess an optimal start-up strategy for the urine nitrification

process in a membrane bioreactor (MBR) configuration, focusing on AOB and NOB adaptation

behavior towards high salinities. Two different strategies were implemented to determine the most

optimal start-up strategy for urine nitrification membrane bioreactors. One MBR was operated at a

high, constant electrical conductivity (salinity) of fully hydrolyzed urine (60 – 70 mS cm-1) while the

other MBR was subjected to a gradual increasing conductivity starting from the normal working

salinity of the inoculum sludge (5 mS cm-1). The first MBR completely lost nitrification activity due to

the salinity shock and after 35 days of adaptation, it was possible to increase the nitrogen loading

rate. When the second MBR reached an electrical conductivity of 15 mS cm-1 (nitrogen loading rate of

250 mg N L-1 d-1), AOB and NOB experienced activity loss and ammonium accumulated. Adaptation

Abstract

xi

lasted for 16 days after which the reactor nitrogen loading rate could be increased to 500 mg N L-1 d-1

in a total of 42 days (100% hydrolyzed urine solution). Full nitrification was achieved in both reactors

with a final reactor electrical conductivity of 45 mS cm-1.

Overall, this thesis shows the potential of biological nitrification for urine stabilization and

demonstrates the high adaptive capabilities of nitrifying sludge. Also, photo-aeration was determined

to have the potential to substitute expensive mechanical aeration in the treatment of source

separated urine. Although, more research is necessary to identify critical inhibiting factors and their

influences on the photobioreactor system and more specifically on the biomass.

Samenvatting

xii

SAMENVATTING

Optimalisatie van duurzame nitrificatie strategiën voor brongescheiden urine.

Urine maakt maar voor 1% deel uit van al het afvalwater dat een waterzuivering bereikt, maar bevat

naar schatting 40% van alle fosfor, 69% van alle stikstof en 60% van alle kalium in het afvalwater. Een

duurzamer alternatief om de conventionele en onefficiënte biologische stikstof en chemische fosfor

verwijdering te vervangen ligt hem in het herwinnen van deze waardevolle nutriënten.

Brongescheiden urine is echter een zeer onstabiele afvalstroom in non steriele condities. Gedurende

collectie en transport wordt de aanwezige organische stikstof gehydrolyseerd en dit zorgt voor een

stijging in de pH met als gevolg ammoniak vervluchtiging. Deze vervluchtinging resulteert in

ongewenste stikstof verliezen en geurhinder. Biologsiche oxidiatie van de aanwezig hoeveelheid

stikstof naar nitraat is een mogelijke oplossing om dit probleem aan te pakken en om de

brongescheiden urine te stabiliseren.

Het eerste deel van dit onderzoek focust op de beluchting die de conventionele biologische

nitrificatie vereist. De kosten voor deze mechanische beluchting beslaan tussen 45 en 75% van het

totale energie verbruik in een waterzuiverings installatie. Een alternatief is de in-situ productie van

zuurstof door middel van photosynthetisch actieve organismen: foto-aeratie. Een eerste doel van

deze thesis bestaat uit de ontwikkeling van een consortium van microalgen en nitrificerende

bacteriën waarin zuurstof, nodig voor nitrificatie, geproduceerd wordt door de microalgen. Dit

proces werd onderzocht in een fotobioreactor. Nitrificatie ging van start na 55 operatie dagen en de

maximaal behaalde nitrificatie efficiëntie was 69% (totale influent stikstof concentratie die omgezet

was naar nitraat) terwijl de stikstofbelading 90 mg N L-1 d-1 was (met een 12% verdunning van

synthetische urine). De bereikte COD verwijderingsefficiënties ware gesitueerd tussen 44 en 83%.

De hoge stikstof, zout en vrije ammoniak in 100% gehydrolyseerde urine limiteerd de activiteit van

ongeadapteerde ammonium oxiderende bacteriën (AOB) en nitriet oxiderende bacteriën (NOB). Het

tweede doel van deze thesis bestaat er dan ook uit om een optimale opstart strategie te bepalen

voor een urine nitrificerende membraan bioreactor (MBR), met het oog op de adaptatie van AOB en

NOB aan de hoge zoutconcentraties die urine met zich meebrengt. Twee verschillende opstart

strategiën werden geïmplementeerd met behulp van twee MBR’s. De eerste MBR opereerde aan de

hoge, constante elektrische conductiviteit (zout concentratie) van volledig gehydrolyzeerde urine (60

– 70 mS cm-1) terwijl de tweede MBR onderworpen was aan gradueel toenemende conductiviteit,

startende vanaf de gebruikelijke conductiviteit van het nitrificerend slib waarmee geïnoculeerd werd

(5 mS cm-1). De eerste MBR verloor zijn volledige nitrificerende activiteit door de hoge initiële zout

concentratie en pas na 35 dagen van adaptatie kon de stikstofbelasting verhoogd worden. Toen de

tweede MBR een conductiviteit van 15 mS cm-1 bereikte (stikstofbelasting van 250 mg N L-1 d-1) werd

een daling in de nitrificatie activiteit opgemerkt aan de hand van ammonium accumulatie in de

reactor. De aanpassingsperiode duurde 16 dagen waarna de stikstofbelasting verhoogd kon worden

tot 500 mg N L-1 d-1 (100% gehydrolyseerde urine) in een periode van 42 dagen. Volledige nitrificatie

werd bereikt in beide membraan bioreactoren met een finale electrische conductiviteit van 45 mS

cm-1.

Deze thesis toont het potentieel aan van biologische nitrificatie voor stabilisatie van brongescheiden

urine en de adaptatie mogelijkheden van nitrificerend slib. Foto-aeratie bleek op laboschaal

potentieel te hebben om, in het behandelen van brongescheiden urine, de dure mechanische aeratie

te vervangen. Om dit process van foto-aeratie te optimaliseren moet er in de toekomst onderzoek

gebeuren naar het effect van de inhiberende factoren op de biomassa en naar betere operationele

strategiën.

PART 1 Literature study

1

PART 1: LITERATURE STUDY

1. INTRODUCTION

1.1. WHY NUTRIENT RECOVERY?

Although almost 80% of atmospheric volume is nitrogen gas, the availability of nitrogen is frequently

the limiting factor both in the production of crops and in human growth. This disparity arises because

virtually all atmospheric nitrogen takes the form of the stable, non-reactive molecule (N2), which

must be split into its two constituent atoms before they can be incorporated into compounds

involved in just about every vital transformation of living matter (Smil, 1991). Natural nitrogen

fixation (transformation of inert N2 to reactive NH3) is carried out by some bacteria (e.g. Azotobacter,

Klebsiella, cyanobacteria, Rhizobium) using the nitrogenase enzyme. This natural process of nitrogen

fixation was not sufficient to sustain the increasing world population. The invention and

commercialization of the Haber-Bosch process made it possible to synthesize ammonia as a source of

active nitrogen for the production of synthetic fertilizers, polymers,... In this process a hydrogen-

nitrogen mixture reacts on an iron catalyst at elevated temperature in the range of 400 – 550°C and

at operating pressures above 100 bar.

This is an exothermic reaction as the change in heat (ΔH) is negative, implying that heat is given off

during the reaction. To achieve the highest possible yield of ammonia low temperatures and high

atmospheric pressure should be used. In practice on the contrary, high temperatures and relative

low pressures are used because high temperatures are easier and cheaper to maintain than high

pressures. An iron catalyst is used to confine losses in ammonia yield. Given the high fuel and

electricity requirements of 35 MJ kg N-1 for this process (Mudahar and Hignett, 1985), it is not

surprising that energy and environmental implications dominate the concerns about the growing

dependence on nitrogen fertilizers (Smil, 1991). This reliance is illustrated in Table 1 (Vance, 2001).

We see that the use of nitrogen fertilizers increased almost ten folds between 1960 and 2000 and is

projected to increase further by 30 Tg yr-1 by 2030 with the increasing world population.

Next to nitrogen also phosphorus is an important macronutrient. P-containing fertilizers find their

application in the enhancing of soil P availability and stimulating crop yields. The use of phosphorus

fertilizer is projected to increase by 20 Tg yr-1 by 2030 (Tabel 1) (Vance, 2001). Phosphorus sources

are restricted and by some estimates, the world’s phosphate rock reserves could be depleted by

2050 (Smil, 2000; Vance et al., 2003).


2

Table 1. Agriculture production and resource use in the recent past to the near future (Vance, 2001).

Item 1960 2000 2030 – 2040

Food production (Mt) 1.8 x 109 3.5 x 109 5.5 x 109

Population (billions) 3 6 8 (maybe 10)

Irrigated land (% of arable) 10 18 20

Cultivated land (hectares) 1.3 x 109 1.5 x 109 1.8 x 109

Water-stressed countries 20 28 52

N fertilizer use (Tg) 10 88 120

P fertilizer use (Tg) 9 40 55 – 60 Data derived from United Nations Food and Agriculture Organization, Waggoner (1994), Bumb

and Baanante (1996), Rosegrant (1997), Dyson (1999), Tilman et al. (2001), and Postel (2001).

Mt, Metric tons; Tg, 1012 g or million metric tons.

At the end of the food chain the nitrogen and phosphorus nutrients end up in the wastewater.

Conventional municipal wastewater treatment plants treat this wastewater so it can be discharged

safely. This implies the discharge without causing eutrophication in natural fresh and salt water

ecosystems. To remove nutrients, the current focus is one on nutrient removal instead of recovery,

with most conventional removal processes being energy intensive and inefficient. This is partly due

to the fact that wastewater is highly diluted and valuable nutrients are spilled. Because of this high

dilution there is a high need for energy to remove or recover nutrients like phosphorus and nitrogen

(Zeeman and Kujawa-Roeleveld, 2011; Zeeman et al., 2008).

Presently, nitrogen compounds are removed from wastewater by a variety of physicochemical and

biological processes. Biological nitrogen removal in the form of nitrification/denitrification and partial

nitritation/anammox (anoxic ammonium oxidation) processes has been widely adopted in favor of

the physicochemical processes. Benefits of the process are the high potential removal of nutrients,

high process stability and reliability, relatively easy process control, low area requirement and

moderate cost. A disadvantage could be that in this process nitrogen gets lost in its inactive form,

nitrogen gas, so that the energy consuming Haber-bosh process has to be used to regain this

nitrogen from the atmosphere (Ahn, 2006; Van Hulle et al., 2010). The primary energy consumption

for the conventional nitrification/denitrification process is 45 MJ kg–1 N (and 109 MJ kg–1 N for

Nitrification/denitrification with methanol as substrate), compared with 37 - 45 MJ kg–1 active

nitrogen produced with the Haber-Bosh process (Maurer et al., 2003). Depending on a life-cycle

analysis of both, recover process and recapture process, it has to be determined which approach is

economically most feasible.

The widespread use of chemical precipitation to remove phosphorous in wastewater treatment

plants results in phosphorus bound as a metal salt within the wasted sludge. Mostly, the sludge is

biologically digested, de-watered and incinerated to reduce the volume and to remove organic

contaminants. But the remaining ash is still rich in heavy metals and the solubility of the aluminum

and iron phosphates is too low to be used directly as a fertilizer (Kuntke et al., 2012). This sludge can

also be used in concrete or it can be discharged to for instance landfills, resulting in a loss of the

phosphorus nutrients. As an alternative to chemical phosphorus precipitation, enhanced biological


3

phosphorus (bio-P) removal (EBPR) has received increased attention. In an anaerobic tank prior to

the aeration tank, polyphosphate-accumulating organisms (PAO) are selectively enriched in the

bacterial community within the activated sludge. These bacteria accumulate polyphosphate within

their cells and the removal of phosphorus is said to be enhanced. Biomass containing biologically

removed phosphorus can be removed by sedimentation and used as fertilizer (Comeau et al., 1986).

A better alternative to replace this inefficient biological nitrogen and chemical phosphorous removal

is the recovery of these valuable nutrients. The physicochemical nitrogen recovery techniques are

difficult to implement in the conventional wastewater treatment because of the diluted

characteristics of domestic wastewater (total nitrogen and phosphorus concentration of respectively

less than 100 mg N L−1 and less than 20 mg P L-1 (Van Hulle et al., 2010)). To tackle this problem it is

interesting to treat streams that are source separated and highly concentrated in nutrients. Examples

of such concentrated streams are source separated urine and effluent of sludge digesters (originated

from concentrated wastewater streams produced in food and agro-industry).

1.2. URINE

1.2.1. COMPOSITION AND PROPERTIES

This human waste stream accounts for only 1% of the total wastewater volume, while it contains

approximately 40% of the phosphorus load, 69% of the nitrogen load, and 60% of the potassium load

that arrives in a municipal wastewater treatment plant (Zeeman and Kujawa-Roeleveld, 2011;

Zeeman et al., 2008). The composition of human urine varies from person to person and from region

to region depending on feeding habits, the amount of drinking water consumed, physical activities,

body size, and environmental factors.

Source separated urine can change in chemical composition during storage. Urea (CO(NH2)2)

becomes hydrolyzed to ammonia and bicarbonate by bacterial urease (urea amidohydrolase)

following equation (1). The catalyst, urease, is produced by a variety of eukaryotic and prokaryotic

organisms but bacteria are the most abundant (many but not all AOB (Koper et al., 2004)). Next to

urease also urea amidolyase is known to hydrolyse urea. This enzyme is known to be produced by

several yeasts and algae (Strope et al., 2011).

( )

(1)

This hydrolysis results in a rise in pH and subsequently enhances precipitation of P compounds like

calcium phosphate, struvite, and calcite. The total result is that 90% of total nitrogen is present as

ammonia or ammonium (mostly ammonia at pH of 9), the pH is close to 9, and at least 30% of

phosphorus is precipitated (Mobley et al., 1995; Udert et al., 2003b).Table 2 summarizes gives

nutrient concentrations in fresh and hydrolyzed human urine (Udert et al., 2006).

Source separated urine can be collected by using source-separated toilets and water free urinals.

Other waste streams rich in nutrients, and in special nitrogen, are anaerobic digester effluents,


4

landfill leachate, specific industrial wastewaters and source-separated domestic wastewater, i.e.

black water digestate (Van Hulle et al., 2010; Vlaeminck et al., 2012).

Table 2. Composition of urine (Udert et al., 2003b).

Component (unit) Fresh urine Hydrolyzed urine

Total nitrogen (mgN L−1) 9200 9200

Total ammonia (mgN L-1) 480 8100

Ammonia NH3 (mgN L-1) 0.3 2700

Urea (mgN L-1) 20 0

Total phosphate (mgP L−1) 740 540

COD (mgO2 L−1) 10000 10000

Calcium (mg L−1) 190 0

Magnesium (mg L−1) 100 0

Potassium (mg L−1) 2200 2200

Total carbonate (mgC L−1) 0 3200

Sulphate (mgSO4 L−1) 1500 1500

Chloride (mg L−1) 3800 3800

Sodium (mg L−1) 2600 2600

Alkalinity (mM) 22 490

pH 6.2 9.1

1.2.2. OPPORTUNITIES FOR NUTRIENT RECOVERY

The separation of wastewater streams and their specific treatment in decentralized systems is a

promising approach. The treatment of source separated urine thereby offers many advantages. Next

to the possibility for more efficient nutrient recycling, it also gives the opportunity of removing

organic micro pollutants originating from the human metabolism (Escher et al., 2006) and offers new

ways of more efficient wastewater management when applied in the rapidly expanding and water-

scarce cities of emerging countries (Huang et al., 2006; Medilanski et al., 2006). The direct and

straightforward application of urine as fertilizer has been a common practice in many rural societies

around the world (Nepal, countries in West Africa,…), but this technique is unsuitable for modern

cities: the high water content makes the transport costly, ammonia volatilization from stored urine is

unpleasant and causes high nitrogen losses, and pathogens pose a health risk to farmers and

consumers of agricultural products (Udert et al., 2006). On top of that, the high salt concentration of

urine causes an increase in salinity of the soil. At present the EU does not include human urine on the

list for approved fertilizers in organic farming which makes its use difficult (Schönning, 2001).

Maurer et al. (2006) discussed the available technologies for treatment of source-separated urine.

They defined seven main purposes of a treatment unit: volume reduction, urine stabilization,

hygienisation, removal of micropollutants, biological nutrient removal, P-recovery and N-recovery.


5

Ammonia stripping under vacuum (0.4 bar, 40°C) from stored urine is a possibility to recover

nitrogen, in which afterwards the gas stream is physically absorbed in water at a pressure of 5 bar

and 20°C. The resulting product contains 10% ammonia and is unstable at normal pressure (Behrendt

et al., 2002). This technology requires a large energy and chemical input (Maurer et al., 2003).

Basakcilardan-Kabakci et al. (2007) reported that in their absorption unit, averagely 92% of ammonia

was recovered as ammonium sulfate, which is a well-known synthetic fertilizer. A variant of this

method is steam stripping which uses steam at 6 bar and 160°C. The energy demand is similar but

the complexity of steam production makes this technology less suitable for small decentralized

reactors (Tettenborn et al., 2007). A more practical method is air stripping which avoid the issues

accompanying the complexity of steam production.

To recover nitrogen, an ion exchanger with a high affinity for ammonium can be used. An example is

the naturally occurring zeolite, clinoptilolite. Ion exchange offers a number of advantages including

the ability to handle shock loadings and the ability to operate over a wider range of temperatures

(Jorgensen and Weatherley, 2003). Up to 97% of the ammonium in stored urine could be transferred

onto clinoptilolite through ion exchange and about 88% could be recovered subsequently from

exhausted clinoptilolite (Beler-Baykal et al., 2011). It has to be considered that when other

monovalent cations (for example Na+) are present, they can also claim cation exchange capacity.

Also isobutylaldehyde-diurea (IBDU) precipitation in fresh urine recovers nitrogen. In fresh, non-

hydrolyzed urine, nitrogen is mainly present in the form of urea. Urea forms a complex with

isobutyraldehyde (IBU), resulting in the precipitation of IBDU, a commercially available slow-release

fertilizer. Although Maurer et al. (2006) showed that this production is not feasible at the urea

concentration of approximately 1% in urine. A pre-concentration step is required. Until know, this

technique was only investigated in laboratory experiments (Behrendt et al., 2002).

Recently, Kuntke et al. (2012) proposed an electrochemical method based on a microbial fuel cell for

energy production combined with nitrogen recovery from urine. The applied microbial fuel cell used

a gas diffusion cathode. The ammonium transport to the cathode occurred due to migration of

ammonium and diffusion of ammonia. In the cathode chamber ionic ammonium was converted to

volatile ammonia due to the high pH. Ammonia was recovered from the liquid-gas boundary via

volatilization and subsequent absorption into an acid solution. An ammonium recovery rate of 3.29 g

N d-1 m-² (vs. membrane surface area) was achieved at a current density of 0.50 A m-². This technique

is still in its research phase.

A biological option to recover nutrients from urine is the cultivation of microalgae. Microalgae take

up and assimilate nutrients in their biomass. Chlorella species, for example, consist of 6–8% nitrogen

and 1–2% phosphorus on a dry weight basis (Oh-Hama and Miyachi, 1988). The microalgae biomass

can be harvested and used as a slow-release bio fertilizer or as a source for the production of high-

value biochemicals and biofuels (Tuantet et al., 2013). But large-scale microalgae production for

biofuels is not yet economically feasible (Kovacevic and Wesseler, 2010).

Udert and Wachter (2012) used biological nitrification with consecutive distillation of the effluent to

concentrate and recover nutrients from urine with a stable and efficient process. The solid residue

contains high amounts of nutrients, such as ammonium nitrate, potassium, sulfate and phosphate,


6

which makes it an interesting product for use in agriculture. A disadvantage was the energy

consumption. A very simple nitrification/distillation setup needs about four to five times the primary

energy that is required to remove the same amount of nitrogen and phosphorus in a conventional

WWTP and produce equivalent amounts of synthetic phosphorus and nitrogen fertilizers.

An example of such a nitrogen fertilizer is ammonium nitrate (NH4NO3). This is a white crystalline

solid that is commonly used in agriculture as a high-nitrogen fertilizer with the NPK rating of 34-0-0

(34% nitrogen). It is more stable than the urea in urine and does not lose nitrogen to the atmosphere

by volatilization. Ammonium nitrate is a quick release fertilizer which means that it is a water soluble

chemical that once applied is readily available to the plant. One of the disadvantages is the high cost,

in comparison with urea, due to production costs and regulations (due to the popularity of

ammonium nitrate in improvised explosives). The ammonia, necessary for the ammonium nitrate

production, is produced with the Haber-Bosch process, with overall reaction being

. This process consumes a lot of energy (the global average energy requirement is at present

somewhere between 40 and 45 GJ ton-1 ammonia (Smil, 1991)). The hydrogen originates from

natural gas, and the nitrogen from air. This use of natural gas is the main cause for the production

related greenhouse gas emissions (Ahlgren et al., 2008).

Another option to produce ammonium nitrate could be the precipitation in partially nitrified urine.

Without pH adjustment only about 50.0% NH4+-N in urine can be converted to nitrate and NH4NO3

can be formed (Feng et al., 2008). Chen (2009) concludes the same, the alkalinity in urine provides

only 41.0% of the alkalinity required for full urine nitrification.

Phosphate recovery is mostly done by the production of magnesium ammonium phosphate

( ), also known as struvite. Next to phosphorus a small fraction of the nitrogen is

recovered which can be stripped. The pH of urine after urea hydrolysis is sufficiently high (pH ≥ 9) to

trigger the precipitation. The precipitation reaction is further triggered by the addition of ,

( ) , or bittern (the magnesium-rich brine from table-salt production) and the struvite can

be used as a synthetic slow-release fertilizer (Udert et al., 2003b).

Flexibility in urine treatment is well illustrated by the many different treatment options discussed,

but there are also many challenges in connection with source separation of urine. Some of them are

the unstable behavior of ammonia and the large amount of urine which needs to be transported and

stored (Basakcilardan-Kabakci et al., 2007). In non-sterile conditions, urea becomes hydrolyzed to

ammonia by bacterial urease. The subsequent pH rise can easily result in ammonia volatilization

which is unwanted because it causes odour problems and more importantly because it means

nitrogen loss. To stabilize urine, biological oxidation of all nitrogen present to ammonium is an

option. Moreover, the nitrification of urine makes it possible to recover water with membranes,

which is not possible with fresh urine because ammonia passes through the pores.


7

2. NITRIFICATION

2.1. GENERAL PRINCIPLES

Nitrification is the sequential biological oxidation of ammonia to nitrate. The process occurs in the

natural environment and is applied in the activated sludge process for nitrogen removal. The first

step of the process is the conversion of ammonia to nitrite according to equation (2), which is

normally done by AOB (Ammonia Oxidizing Bacteria). This reaction is called nitritation.

(2)

The nitritation reaction consists of two sequential oxidation steps. First ammonia (NH3) is oxidized to

hydroxylamine (NH2OH) with the enzyme ammonia monooxygenase (AMO). The second step involves

the oxidation of hydroxylamine to nitrite (NO2−) with the enzyme hydroxylamine oxidoreductase

(HAO) (Kowalchuk and Stephen, 2001). During the oxidation of ammonia to nitrite, hydrogen ions are

released and thus decrease the pH in case no sufficient alkalinity is present.

The second step, the nitratation, is the conversion of nitrite to nitrate by NOB (Nitrite Oxidizing

Bacteria) according to equation (3).

(3)

The complete nitrification can be described by equation (4).

(4)

Nitrification is mediated by nitrifying bacteria. These AOB and NOB are aerobic

chemolithoautotrophic microorganisms. Some heterotrophic bacteria can also oxidize ammonia to

nitrate, but this is only a very small contribution to the overall ammonia oxidation (Van Hulle et al.,

2010). The nitrifying bacteria obtain carbon from dissolved CO2 and energy for growth from oxidizing

ammonia and nitrite. One mole of carbon dioxide to be assimilated into nitrifying bacteria requires

approximately 30 moles of ammonium ions or 100 moles of nitrite ions to be oxidized (Chen, 2009;

Gerardi, 2002).

AOB, which are responsible for the ammonia oxidation, have 25 cultured species, among which

Nitrosomonas is the most extensively studied genus. Next the bacteria, also ammonia-oxidizing

archaea (AOA) can oxidize ammonia. Nitrosopumilus maritimus and Nitrososphaera viennensis, have

been isolated and described (Martens-Habbena et al., 2009) amongst others. NOB which are

responsible for the nitrite oxidation have many cultured species of which numerous described in the

last decade (Koops and Pommerening-Roser, 2001). They differ in ecophysiological requirements.

Nitrospira were found to be the dominant nitrite oxidizers in both enriched and full scale nitrifying

systems (Abeliovich, 2006).

AOB and NOB obtain carbon from dissolved . The pH range of the 2 reactions is usually in the

range of 5.5 to 8.3, which makes bicarbonate the main alkali species according to equation (5).


8

(5)

Assuming the empirical formula of bacterial cells to be C5H7O2N, the full stoichiometry’s of

ammonium and nitrite oxidation can be described by equations (6-7) (Wiesmann and Libra, 1999).

(6)

(7)

The equilibrium of the carbonate system in water plays an important role in nitrification and depends

strongly on the pH. This carbonate system is usually the pH buffer available in wastewater and

neutralizes the production of protons. The equilibrium is given in equation (8).

( ) ( )

(8)

The overall synthesis and oxidation reaction of nitrifying bacteria is given by equation (9).

(9)

The total oxygen consumption is 4.27 mg O2 for nitrification of 1 mg NH4+-N to NO3

--N. This equation

also indicates that nitrification of 1 mg ammonium requires 7.07 mg of alkalinity.

2.2. PROCESS CONDITIONS FOR NITRIFICATION

The nitrification process in wastewater treatment plants is restricted to only a few genera of

bacteria. This fact and their slow growth make the nitrification process very susceptible to inhibition.

This vulnerability of the biological nitrification process makes it important to understand the biology

and requirements of AOB and NOB (Grunditz and Dalhammar, 2001).

2.2.1. FREE AMMONIA AND FREE NITROUS ACID INHIBITION

Ammonia (NH3) rather than ammonium ( ) is the preferred substrate for oxidation (Suzuki et al.,

1974). Ammonia is nonetheless toxic and nitrification inhibition occurs when concentrations become

too high (Table 3). Gerardi (2002) showed that in a temperature range of 10 – 20 °C and a pH range

of 7 to 8.5, about 95 % of the reduced form of nitrogen is . Equation (10) gives the equilibrium.

(10)

Also nitrite is toxic and can cause nitrification inhibition. The nitritation and nitratation reactions can

proceed at different velocities. If the nitritation is faster than the nitratation, there will be an

accumulation of which is in equilibrium with HNO2. The build-up of HNO2, according to equation

(11), can be problematic since it is poisonous to both AOB and NOB.

(11)

Concerning inhibition it can be stated that NH3 is the main inhibitor of nitrification at high pH (>8),

whereas HNO2 is the main inhibitor at low pH (<7.5). Table 3 shows the inhibition concentrations for

NH3 and HNO2 towards AOB and NOB (Anthonisen et al., 1976).


9

Table 3. Inhibition of AOB and NOB by free ammonia (FA) and free nitrous acid (FNA)

(Anthonisen et al., 1976).

Component AOB NOB

FA (mg N/L) 8 – 120 0.08 – 0.82

FNA (mg N/L) 0.2 – 2.8 0.06 – 0.83

2.2.2. PH

The nitrifiers require an alkaline environment with an optimum pH for Nitrosomonas and Nitrobacter

between 7.5 and 8. The preference of AOB for slightly alkaline environments results from the fact

that these organisms use NH3 as substrate while at certain pH values NH3 and HNO2 can exhibit

inhibitory effects as stated above. Below pH 7.0, nitrification rate will decrease since carbon

limitation due to CO2 stripping will occur (Wett and Rauch, 2003). Below a pH of 6.0 the nitrification

stops. For ammonia oxidizers, this is due to increasing ionization of (Figure 1), which enters the

cell by diffusion, to , which enters the cell by active transport.

Figure 1. The dependence of the NH4+/NH3 ratio as a function of pH (Clegg and Whitfield, 1995).

Suthersand and Ganczarczyk (1986) found that AOB are more sensitive to pH changes than nitrite

oxidizers, suffering irreversible activity loss during pH shocks. Willke and Vorlop (1996) found that

NOB have a great resistance to pH changes. Their activity stays unaffected at pH values between 4.5

and 10.

Philips et al. (2002) demonstrated with growth experiments, in which both AOB and NOB were

provided with optimal conditions, that AOB can grow faster than NOB at a temperature of 28°C. The

minimum doubling time of AOB and NOB is respectively 7-8 hours and 10-13 hours. The reason for

this difference in growth velocity is that the energy delivered by the oxidation of to

(274.91 kJ mol-1) is much higher than the energy delivered by the oxidation of to (74.16 kJ

mol-1).


10

2.2.3. OXYGEN

AOB and NOB, known as ‘nitrifiers’, are strict aerobes. Peng and Zhu (2006) stated that when the

oxygen level drops lower then 1.0 - 1.5 mg O2 L-1 the nitritation reaction still takes place but the

nitratation reaction is inhibited depending on applied temperature. Kampschreur et al. (2007)

showed that, in a nitrifying reactor treating wastewater, denitrification by AOB is the main nitrous-

oxide-producing pathway and the emission was increased with nitrite presence and oxygen

limitation. This nitrous oxide has a negative effect on the greenhouse effect. According to Hunik et al.

(1994), the half-saturation constant (Ko) for dissolved oxygen is 0.16 mg O2 L-1 for the ammonium

oxidizer Nitrosomonas europaea and 0.54 mg O2 L-1 for the nitrite oxidizer Nitrobacter agilis,

respectively. Picioreanu et al. (1997) however, states dissolved oxygen half-saturation coefficients for

AOB and NOB of 0.2–0.4 mg L-1 and 1.2–1.5 mg L-1, respectively. Different half-saturation constants

are found in literature because the constant is dependent on the biomass density, the floc size, the

mixing intensity, the rate of diffusion of oxygen in the floc and the exact nitrifying bacteria present

(Van Hulle et al., 2010).

2.2.4. TEMPERATURE

Temperature is an important parameter in the nitrification process, but the exact influence is difficult

to determine because of its interaction with mass transfer, chemical equilibria and bacterial growth

rate. A temperature rise creates several opposite effects. First, a higher temperature means

increased NH3-level inhibition but decreased FNA levels, second, a higher temperature will increase

the activity of the organisms according to the Arrhenius principle but there will also be more decay.

On top of that, higher oxygen diffusivity will be present, but also lower oxygen solubility follows from

the Arrhenius principle. The increased temperature activity only holds up to a certain critical

temperature above which biological activity decreases again. Experiments with pure cultures gave an

optimal temperature of 35 °C for ammonium oxidizers and 38 °C for nitrite oxidizers (Grunditz and

Dalhammar, 2001). At temperatures of 40 °C and higher, nitrification rates fall to near zero. At

temperatures below 20 °C, nitrification proceeds but at a slower rate (Bothe et al., 2007). Schmidt et

al. (2003) give values for activation energy of ammonium and nitrite oxidizers of 70 kJ mol-1 and 44 kJ

mol-1, respectively. This indicates that the activity of ammonium oxidizers will increase faster than

the activity of nitrite oxidizers. Because cooling costs for warm wastewaters can run high, Courtens

et al. (2014) showed that mesophilic nitrifying sludge can gradually adapt to 42.5 and 47.5 °C for

nitritation and nitratation, respectively.

2.2.5. SALINITY

Ye et al. (2009) stated the higher sensitivity of NOB to high salt concentrations compared with AOB in

aerobic biological treatment. The number of NOB decreased strongly when the salinity was above

10 g L-1 and the presence of NOB was less than 1% when salinity was higher than 20 g L-1 while the

viability of AOB was only half of the original condition. Di Bella et al. (2013) reported decreasing

ammonium removal efficiencies from 99% to 80% inside a membrane bioreactor, as consequence of

a salt concentration of 10 g NaCl L-1 (or 16.95 mS cm-1) and at 20 g NaCl L-1, nitrification activity

reached zero. Johir et al. (2013) concluded that the salt tolerance ability of fresh nitrifiers could be up


11

to 15 g L-1 of NaCl. Courtens et al. (2014) showed that for a mesophilic nitrifying community, more

efficient temperature transition and eventually higher thermotolerance can be reached through

addition of NaCl at 5 to 7.5 g L−1

2.2.6. LIGHT

Light is inhibiting both ammonium oxidizers and nitrite oxidizers since cytochrome c is oxidized by

light in the presence of oxygen (Van Hulle et al., 2010). Alleman et al. (1987) researched the impact

of light exposure on an enriched Nitrosomonas culture and concluded that the range of 410 – 415 nm

particularly appears to be responsible for Nitrosomonas inhibition. Also, full recovery of the bacteria

was observed after 6 – 10 h of dark exposure with ammonium presence. No records were found for a

long period of illumination or potential adaptation to light inhibiting conditions.

2.3. NITRIFICATION OF URINE

Urine nitrification for stabilization has been extensively studied in literature (Table 4). Different

configurations with different concentrations of synthetic and real human urine have been reported.

Overall, the nitrification is a difficult process because of the possible free ammonia, the unstable pH

and high salt concentration. The adaptation of the culture after the start-up phase can help to

prevent these problems.


12

Table 4. Comparison of reactor tests for urine nitrification.

Suspended biomass Attached biomass

Chen, 2009 Sun et al., 2012 Udert et al., 2003 Udert et al., 2003 Sun et al., 2012 Udert et al., 2003 Feng et al., 2008

Reactor technology

SBR SBR SBR CSTR MBR(3) MMBR Packed bed bioreactor

Dilution % artificial urine

N/A 15 (no organics) N/A N/A 19 (no organics)

N/A 12.5

Dilution % real human urine

10-29 28-47 13-42 137(2) 28 16-87 10

N oxidation rate (gN L-1 d-1)

1.05 0.40-0.75 0.3-1.3 0.8 0.25 0.380 0.044

pH 7.6 6.4-6.5(1) 6.0-8.8(4) 6.9 6.3 7.0- 7.8 6.65

T (°C) 25 25-35 24.5 30 35 25.3 27

DO (mg O2 L-1) 1.5 2 2.0-4.5 2.0-4.5 3 3.0-5.2 4.29

Inoculum Undefined Undefined Undefined Undefined Undefined Undefined Undefined

Effluent 100% NO3- 50% NO2

-

50% NH4+

50% NO2-(5)

50% NH4+

50% NO2-(5)

50% NH4+

50% NO2-

50% NH4+

50% NO3-

50% NH4+

95% NO3-

5% NH4+

(1) With some higher outliers (7.6 – 8.9)

(2) Electrical conductivity corresponded more or less to undiluted urine, yet the N concentration was artificially altered through NH4HCO3 addition

(3) Submerged hollow-fibre membranes

(4) Range over operational cycle

(5) Intended outcome, for partial nitritation/anammox or chemical nitrite oxidation experiments

(N/A) Not applicable


13

2.4. MEMBRANE BIOREACTOR

Nutrient recovery techniques require a high effluent quality and thus a high activity of the nitrifying

sludge. This is possible in a compact manner using a membrane bioreactor configuration. The basic

membrane bioreactor consists of two processing steps. First there is a bioreactor in which aerobic

bacteria digest COD and nitrogen in the presence of dissolved oxygen. Secondly, a membrane

module separates the treated water from the suspension of organic matter and bacteria.

The membrane unit replaces the settling stage of the conventional activated sludge system

processing in a much smaller space, which makes it much more attractive for construction in

developed urban areas. The excess solids created by the oxidation process can then be easily

removed for subsequent treatment. It is a continuous process, and one that is quite easily controlled.

The quality of rejected water is independent of the variations of sludge settling velocity. This leads to

the major advantage that it can operate at a much higher solids concentration than that of a

conventional activated sludge plant. The MBR plant can work effectively at MLSS (mixed liquor

suspended solids) concentrations typically in the range 8 to 12 g L-1 (Yamamoto et al. (1989) utilized

even 30 g L-1), whereas conventional activated sludge plants work at about 2 to 3 g L-1, because of the

limitations on settling. This high sludge concentration capability enables an MBR system to deal

effectively with concentrated waste streams like urine. The membrane bioreactor is characterized by

a complete retention of the biomass inside the bioreactor, which controls and increases the sludge

retention time (SRT) independently from the hydraulic retention time (HRT). High SRTs makes it

possible to increase the sludge concentration and the applied organic load, thereby increasing the

pollutant degradation (Marrot et al., 2004). High sludge retention time allows the systems to keep

sufficient amounts of slow-growing microbes, such as AOB and NOB (Chen and LaPara, 2008).

Additionally, the high MLSS concentrations in the MBR which result in low F/M (feed/biomass) ratios

are beneficial for nitrifiers (Liu et al., 2005a). These conditions decrease the selection pressure on the

nitrifiers and support their growth under stress conditions, such as low temperature or inhibition

(Liebig et al., 2001). Although the process has many advantages, there are disadvantages of

accumulation of inert compounds and reduction of biomass viability with operation time (Hasar et

al., 2004).

There are two types of configurations for the membrane array (Figure 2): the membranes can be

placed either outside or inside the bioreactor. For the external or side stream configuration, the two

units are set up to run in succession and the mixed liquor is filtered under pressure in a specific

membrane module. For the submerged or immersed configuration, the filtration is carried out in the

aeration basin by suction removal of the effluent. The submerged configuration appears to be more

economical based on energy consumption. No recycle pump is needed since aeration generates a

tangential liquid flow in the vicinity of the membranes, and it operates at lower values of TMP and

tangential velocities (Marrot et al., 2004).


14

Figure 2. Left: External loop process. Right: Submerged membrane process (Carstensen et al., 2012).

2.5. AEROBIC SLUDGE GRANULATION

Granular sludge is the dense accumulation of high amounts of active biomass due to microbial self-

aggregation (Chen, 2009). Anaerobic granulation has its application in up-flow anaerobic sludge

blanket (UASB) reactors, but more recently, aerobic granules have been developed. Aerobic granules

possess high settling velocity, high biomass retention, strong microbial structure and high resistance

to inhibitory and toxic wastes (Qin et al., 2004b; Tay et al., 2001).

Many factors influence the aerobic granule formation, including substrate composition, organic

loading, hydrodynamic shear force, feast-famine regime, feeding strategy, DO, reactor configuration,

SRT, cycle time, settling time, and exchange ratio (Liu and Tay, 2004). Liu et al. (2005b) further

reported that hydraulic selection pressures on SBR’s sludge particles would mainly contribute to the

formation of aerobic granules. Many studies agree that substrate loading rate and settling time is the

most important factor influencing of sludge granulation (Chen et al., 2008; Qin et al., 2004a; Wang et

al., 2004).

The substrate loading rate for the aerobic granules formation can be flexible from 2.5 to 15 kg COD

m-3 d-1 (Chen et al., 2008). Nitrifying sludge granules can also be developed over a wide range of

ammonia nitrogen loading rates (Tay et al., 2006; Tsuneda et al., 2004). To facilitate microbial

aggregation, cells can be made more hydrophobic by being subjected to a periodic feast and famine

regime (periodic starvation). Aggregation could be an effective strategy for cells against starvation

(Bossier and Verstraete, 1996).

Because long settling times enhance the retention of poorly settling sludge flocs, which may compete

with granule-forming bioparticles, a short settling time is required to enhance the aerobic granules

formation in SBR (Liu et al., 2005b; Qin et al., 2004a). Also a high hydrodynamic shear force is in favor

of the formation of compact and dense aerobic granules (Tay et al., 2002).

Divalent ions could accelerate aerobic granules formation. Addition of 100 mg Ca2+ L-1 or 10 mg Mg2

+

L-1 significantly decreased the formation time of aerobic granules (Jiang et al., 2003). Furthermore,

slowly growing bacteria like nitrifying bacteria were also found to facilitate the granules formation


15

and results in dense granules with good cell hydrophobicity (Liu and Tay, 2004). Under a high

substrate N/COD ratio, nitrifying population in the aerobic granules was enriched remarkably.

3. MICROALGAE AND THEIR APPLICATIONS IN WASTEWATER TREATMENT

3.1. MICROALGAE

The term microalgae refers to all algae too small to be seen without the use of a microscope, and

includes both eukaryotic microalgae and phototrophic prokaryotes (for example cyanobacteria).

Microalgae therefore exist in a range from small, unicellular particles to more complex aggregated

multicellular ones. In comparison with higher plants, they do not differentiate and lack specialized

organs such as roots, stems, flowers and leaves (Bosma, 2010). Mostly they are autotrophs that are

capable of photosynthesis and use sunlight, CO2 and nutrients (for example nitrogen and

phosphorus) to grow (Deloof, 2010; Larsdotter, 2006). Many species can also grow mixotrophically

and some even heterotrophically.

3.2. APPLICATION OF MICROALGAE IN A WWTP

Wastewater treatment with microalgae is not a new concept and is developed in the 1950s (Oswald

et al., 1963; Oswald et al., 1957). The nitrogen and phosphorus nutrients are assimilated into

valuable algal biomass using photosynthesis (de la Noüe and De Pauw, 1988). Boelee et al. (2012)

used following stoichiometric reactions for microalgae metabolism, with either nitrate or ammonium

as nitrogen source:

Optimal wastewater N/P ratios for microalgae are in the range of 16:1, known as the Redfield ratio

(Redfield, 1958). According to the ratios most often found in wastewater, nitrogen may be the

limiting nutrient. Urine has an N/P atomic ratio of 13:1. This ratio can increases after urine hydrolysis

because an important amount of phosphate precipitates as struvite or hydroxyapatite (Udert et al.,

2003b).

In conventional wastewater treatment plants phosphorus and nitrogen are largely lost during

treatment. Nitrogen is mainly removed as N2-gas to the atmosphere by dissimilatory nitrification and

denitrification processes. In this perspective, treatment with microalgae offers an interesting

approach to remove nutrients from wastewater. Another opportunity is the possible valorization of

the produced biomass. Wastewater fed microalgae have been used for lab-scale production of oil,

biodiesel, biogas, bio-ethanol, fertilizer, animal feed ingredients and high-value molecules such as

pigments, amino acids and poly-β-hydroxybutyrate (PHB) (Van Den Hende, 2014). Although in most


16

cases their economic feasibility on industrial scale remains challenging to demonstrate under non-

arid and highly populated western areas.

The most important parameters that regulate algal growth are light, pH, nutrient quantity and

quality, turbulence (stirring), salinity and temperature. The optimal parameters are species specific

but provided sufficient light and space are available and the wastewater is non-toxic, treatment with

microalgae is possible (Coutteau, 1996). Bhaya (1996) showed that when ammonium is available, no

alternative nitrogen sources will be assimilated. In addition to ammonium and nitrate, urea

((NH2)2CO) and nitrite (NO2-) can also be used as nitrogen sources. However, the toxicity of nitrite at

higher concentrations makes it less convenient (Becker, 1994).

There are two main cultivation systems: open raceway ponds (stabilization ponds or high rate algal

ponds) and closed photobioreactors (dome, tubular, flat plate systems) (Munoz and Guieysse, 2006).

Table 5 lists advantages and disadvantages for these cultivation systems. Biofilm-based reactors

might be subject of extra limitations like an increased photo-inhibition due to constant exposure to

high light intensities at day and the potential risk of clogging due to biomass growth in case of PBRs

(Munoz et al., 2009).

Table 5. Comparison of open and closed systems for microalgae systems +: advantage; -:

disadvantage (Van Den Hende, 2014).

Characteristic Open system Closed system

Investment and maintenance costs + Low - High

Construction and operation + Easy - Difficult

Biomass densities - Low (high if biofilm) + High

Photosynthetic efficiencies - Low + High

Risk of predator contamination - High + Low

Biomass productivities - Low + High

Illuminated surface:volume ratio - Low + High

Needed land area - High + Low

Toxic accumulation of O2 + Low - High

Overheating of reactor + No problem - High in summer

Difficult to remove biofouling of

reactor wall (if no biofilm reactor)

+ Low - High

To be effective in wastewater treatment, the algal biomass must be efficiently removed. Because

microalgae are very small and usually have slow settling rates, efficient removal is difficult and

expensive in most cases (de la Noue et al., 1992). The harvesting cost can contribute for 20 – 30% of

the total algal biomass production cost, depending on species, cell density, and culture conditions

(Grima et al., 2003). Usually, harvesting includes a solid-liquid separation followed by dewatering and

drying (Abdel-Raouf et al., 2012).


17

3.3. PHOTOSYNTHETIC AERATION

Mechanical aeration accounts for 45 – 75% of the total energy consumption in biological wastewater

treatment plants (Stenstrom and Rosso, 2008). An alternative for mechanical aeration is the in-situ

production of oxygen by photosynthetic organisms, or photoaeration. Photosynthetic aeration is the

process in which photosynthetic organisms produce oxygen for aeration. The process is the basic

working principle of oxidation ponds, also called lagoons or stabilization ponds, which are large,

shallow ponds designed to treat wastewater through the interaction of sunlight, bacteria, and algae.

Algae use energy from the sun and carbon dioxide and inorganic compounds released by bacteria in

water (Oswald et al., 1957). MaB-flocs (micro-algal bacterial flocs) which are aggregations of

microalgae and bacteria are a perfect example of photoaeration. In these flocs, CO2 is captured

within the algae biomass and oxygen is produced in situ. Van Den Hende (2014) tested the outdoor

and year round performance of MaB-flocs for wastewater combined with flue gas treatment. Karya

et al. (2013) showed that a mixed culture of algae and nitrifying micro-organisms can be maintained

in a photo-bioreactor fed with artificial wastewater 50 mg NH4+-N L-1 and achieved 100% nitrification,

without aeration. The algae produced oxygen at a rate of 0.46 kg m-3 day-1 which is higher than for

waste stabilisation ponds (0.01 kg m-3 day-1) and in HRAPs (0.3–0.38 kg m-3 day-1).

3.4. ALGAE ON URINE

The use of diluted urine to cultivate microalgae and cyanobacteria (blue-green algae) is a relatively

new research topic. It was first studied for life support systems for space missions (Tuantet et al.,

2013). Studies showed that it is possible to grow microalgae on highly diluted human urine

(Adamsson, 2000; Feng and Wu, 2006; Yang et al., 2008). Tuantet et al. (2013) was the first to

investigate the growth of microalgae on pure human urine. In addition, they determined the

essential requirements for microalgae growth on different types of urine (fresh, hydrolyzed, male

and female). Tuantet et al. (2013) concluded that it is possible to grow C. sorokiniana on diluted urine

up to 1.4 g NH4+-N L-1 at a pH lower than 8.0 and at slightly reduced growth rates. The study showed

that urine can serve as a rich source of major nutrients for the large-scale sustainable production of

microalgae biomass.


18

4. RESEARCH OBJECTIVES

Source separated urine is an interesting waste stream for nutrient recovery, due to its high

concentrations of nitrogen (6 – 8 g L-1) and phosphorus (0.5 g L-1). This concentrated waste stream is

however very unstable and pH rise due to urea hydrolysis can easily result in ammonia volatilization.

Nitrification of urine is a possible solution to tackle this problem.

A problem in conventional wastewater treatment is the aeration cost for biological nitrification.

Photo-aeration can offer a possibility to reduce these aeration costs for nitrification in conventional

WWTP. A first goal of this research project is to develop and study a consortium of microalgae and

nitrifying bacteria, in which microalgae produce the oxygen in-situ which is used by the nitrifiers to

oxidize ammonium. This process is studied in a bubble column photobioreactor.

The high nitrogen concentration, salinity and free ammonia in urine limit the activity of non-adapted

AOB and NOB. A second goal of this research project is to assess an optimal start-up strategy for the

urine nitrification process in an MBR configuration. An important element hereby is the adaptation of

AOB and NOB to high salinities, which are related to fully hydrolyzed urine.

PART 2 Materials and Methods

19

PART 2: MATERIALS AND METHODS

1. PHOTOBIOREACTOR BUBBLE COLUMN

1.1. REACTOR DESIGN AND SET UP

To design and build a photobioreactor it was important to make some decisions about the

configuration (tubular reactor, panel reactor, closed or open…), the volumetric size of the reactor,

the spatial dimensions, mixing of the reactor content and so on. A plexiglass, gastight, bubble column

photobioreactor with a wide internal diameter of 12.0 cm and a height of 50 cm was the result. It

was operated in sequential batch mode (SBR). A wide column diameter of 12 cm was chosen to

temper algal growth and protect the nitrifying bacteria from light inhibition. The working volume was

4.0 L and the illuminated surface area was 0.16 m². Fluorescent growth lamps (Grolux T5, 24W,

Sylvania) provided a photon flux density of 300 µmol PAR m-2 s-1 which was measured with a light

meter at the outer reactor wall (Fieldscout quantum light meter, USA).

Figure 3 gives a schematic reactor overview and Figure 4 shows the laboratory reactor set up.

Initially CO2 and N2 gas was introduced by means of an air diffuser at the bottom of the reactor. The

gas flow rate was controlled with two gas-flow controllers (Bronkhorst high-tech EL-FLOW select

mass flow meter/controller) and was chosen in that way the volume of flue gas was 1% of the gas

recirculation flow. Four pumps were used: one pump for liquid recirculation (Watson-Marlow 503 S),

one peristaltic pump to dose the influent (ProMinent® DulcoFlex DF2a peristaltic pump), one

peristaltic pump (ProMinent® DulcoFlex DF2a peristaltic pump) to extract effluent and one vacuum

pump (VWR Mini diaphragm vacuum gas pump) for headspace gas recirculation. The recirculation of

the headspace was chosen for practical reasons since a mixing device to keep the reactor content in

suspension, requires a lot of technical skills to install gastight.


20

Figure 3. Reactor Setup: schematic overview of the lab-scale photobioreactor bubble column.

Figure 4. Reactor setup: lab-scale photobioreactor bubble column.

Influent and effluent pump

Flow controller

Vacuum pump

Recirculation pump

pH controller


21

Aeration Settling Effluent

460 minutes

20 minutes

8 minutes

Influent

8 minutes

A pH controller (Prominent® dulcometer; prominent pH electrode dulcotest PHEP-112) was installed

on the recirculation circuit to keep the pH between the desired levels of 6.5 and 7. A DO probe was

used to check the dissolved oxygen level. Excess of produced gas could escape through an

overpressure exhaust channel on top of the reactor. To prevent air from coming in through this

exhaust vent, a water lock was installed. The reactor fluid was kept in suspension only by means of

the gas flow, the liquid recirculation and the gas recirculation. The headspace was monitored for CO2,

N2O and CH4 with a compact GC (Global Analyser Solutions, Breda, The Netherlands). The

temperature was kept at 20 °C in a temperature controlled room. To exclude oxygen leakage,

gastight tubing was used for the gas recirculation (Masterflex norprene). To prevent settling of

biomass, which can cause anaerobic zones with consequently denitrification and nitrogen loss, a

conical bottom with an aeration ring was implemented.

1.2. OPERATION

The initial feed solution was a 10% diluted synthetic urine medium which implied a nitrogen loading

rate of 0.05 g N L-1 d-1. A trace solution was added to provide the necessary nutrients for biomass

growth and the salt concentration was altered until the solution had the conductivity of a 33%

dilution of urine (6.66 mS cm-1).

Since urine contains a high nitrogen and COD concentration, an appropriate feeding pattern was

necessary to avoid possible FA (Free Ammonia) and FNA (Free Nitrous Acid) inhibitory effects.

Instead of dosing the daily influent in one time only, 3 cycles were used during 24 hours to spread

the nitrogen loading over a longer reaction time. The timeline of 1 cycle is illustrated in Figure 5.

Figure 5. Timeline for 1 operation cycle of the PBR.

The duration of one complete cycle was 8 hours in which 7.7 hours were aerated and 0.3 hours were

foreseen for sludge settling. During the last 8 minutes of the settling period and during the first 8

minutes of a new cycle, supernatant (effluent) was extracted and influent was dosed respectively.

During each cycle, 0.2 L of influent was dosed into the reactor and each day 3 complete cycles were

performed, corresponding to an HRT of 6.67 d. The operation principle of this cycle was chosen like

this to implement a feast/famine regime accompanied with short settling times to select for the

fastest settling flocs and wash out of badly settling organisms.

Table 6 gives an overview of the initial and final operation parameters.


22

Table 6. Overview of the initial and final PBR operation parameters.

Operation parameter Value

Working volume (L) 4.0

Q liquid recirculation (L min-1) 0.5

Q gas recirculation (L min-1) 6

pH 6.5 - 7

Temperature (°C) 20

Settling velocity (m h-1) 5

Δt cyclus (h) 8

Nitrogen loading rate (gN L-1 d-1) 0.05 – 0.09*

Q influent/effluent (mL d-1) 333 – 600*

HRT (d) 12 – 6.7*

*final operation parameter

1.3. SYNTHETIC URINE

The artificial urine medium developed by Brooks and Keevil (1997) was used as reactor inlfuent. The

composition of the synthetic urine is given in Table 7. The synthetic urine was stored at 4 °C prior to

be fed to the reactor and was refreshed every week to avoid hydrolysis of the urea. Because the final

influent target was a 33% dilution of this synthetic urine and 100% synthetic urine has an electrical

conductivity of 18.73 mS cm-1, the influent electrical conductivity was altered to 6.66 mS cm-1. Also

trace element solution was added (Kuai and Verstraete, 1998).

Table 7. Composition of synthetic urine.

Component Quantity (g) Concentration (mmol L-1)

Peptone L37 1

Yeast extract 0.005

Lactic acid 0.1 1.1

Citric acid 0.4 2

Sodium bicarbonate 2.1 25

Urea 10 170

Uric acid 0.07 0.4

Creatinine 0.8 7

Calcium chloride.2H2O 0.37 2.5

Sodium Chloride 5.2 90

Iron II sulphate.7H2O 0.0012 0.005

Magnesium sulphate.7H2O 0.49 2

Sodium sulphate.10H2O 3.2 10

Potassium dihydrogen phosphate 0.95 7

Di-potassium hydrogen phosphate 1.2 7

Ammonium chloride 1.3 25


23

1.4. SLUDGE AND ALGAE

The initial inoculum for the reactor comprises of washed and activated HANDS (Higly Active Nitrifying

and Denitrifying Sludge) (Avecom, Belgium), 7 species of microalgae and 1 microalgae mix acquired

from a grassland pond. The 7 microalgae species comprised Chlorella sp., Haematococcus sp.

(CCMP3417), Desmodesmus sp., Ankistrodesmus sp., Pediastrum duplex, Chlorella vulgaris and

Nannochloropsis (CCMP531) (which occurs mostly in marine environments). These cultures were

acquired from the Marine Biology Section MARBIO, Ugent, located at campus De Sterre. The exact

initial TSS concentration was set on 2 mg TSS L-1. The amount of algae, necessary to provide enough

oxygen for the nitrifiers, was calculated.

The algae were cultivated in 0.5 L Erlenmeyer flasks that were placed on a shaker (Inova 2300

platform shaker) in a temperature controlled room at 20°C. The used culture medium was ESAW

(Enriched Seawater Artificial Water) medium made with artificial seawater (Harrison et al., 1980).

The recipe is given in Table 8. The flasks were aerated with compressed air which was humidified by

pumping it through a closed barrel filled with water. Humidification of the air was done to prevent

evaporation from the flasks. The flasks were closed from the environment with cotton wool to

prevent contamination. The algae were exposed all times to light from 7 TL lamps (Philips TL-D 90 De

Lux 36W). Figure 6 illustrates the algae cultivation.

Figure 6. Algae cultivation before reactor inoculation. Left: Shaker in the temperature controlled

room; Right: Erlenmeyer flasks containing the microalgae cultures.


24

Table 8. ESAW culture medium (Harrison et al., 1980).

Component Concentration (g L-1) Concentration (mM)

Solution I - Anhydrous salts

NaCl 20.756 326.7

Na2SO4 3.477 25.0

KCl 0.587 8.03

NaHCO3 0.17 2.067

KBr 0.0845 0.725

H3BO3 0.022 0.372

NaF 0.0027 0.0627

Solution II – Hydrated salts

MgCl2 . 6H2O 9.395 47.18

CaCl2 . 2H2O 1.316 9.134

SrCl2 . 6H2O 0.0214 0.082

Nutrients and trace elements

NaNO3 46.67 549.1

Na2glyceroPO4 6.67 21.8

Na2SiO3 . 9H2O 15.00 105.6

Na2EDTA . 2H2O 3.64 9.81

Fe(NH4)2(SO4)2 . 6H2O 2.34 5.97

FeCl3 . 6H2O 0.16 0.592

MnSO4 . 4H2O 0.54 2.42

ZnSO4 . 7H2O 0.073 0.254

CoSO4 . 7H2O 0.016 0.0569

Na2MoO4 . 2H4O 0.126 0.520

Na2EDTA . 2H2O 1.89 5.05

H3BO3 3.80 61.46

Na2SeO3 0.00173 0.001

Vitamins

Thiamine 0.1 2.97 * 10-1

Vitamin B12 0.002 1.47 * 10-3

Biotin 0.001 4.09 * 10-3


25

2. MEMBRANE BIOREACTORS

2.1. REACTOR SET UP

Two submerged membrane bioreactors (MBR) with the dimensions of 50x20x10 cm were operated in

semi-continuous mode. The working volume was 8.0 L. A commercially available Kubota membrane

was used with a surface area of 0.8 m², a pore diameter of 0.4 µm and a capacity of 1 – 1.5 m³ d-1.

Figure 7 gives a schematic reactor overview and Figure 8 shows the lab-scale reactor set up. Air was

introduced through an air diffuser at the bottom of the reactor to provide enough oxygen and to

keep the biomass in suspension and prevent biofilm formation on the membrane. Air was introduced

at the bottom of the reactor by means of a vacuum pump with a flow rate of 5 L min-1 (VWR Mini

diaphragm vacuum gas pump) and extra air was introduced with the aid of two aquarium pumps

(Rena air 600) and air diffuser stones. Four pumps were used: two peristaltic pumps (Watson Marlow

323 S) to dose the influent and two peristaltic pumps (Prominent 1.6) to extract effluent. A pH

controller (Prominent dulcometer; prominent dulcotest pH electrode PHEP-112) was placed directly

in the reactor fluid to keep the pH between the desired levels of 6.9 and 7.1.

The influent and effluent of MBR 1 and 2 was sampled daily at the same moment during 60 days. The

samples were analyzed for ammonium, nitrite and nitrate. pH and DO were monitored daily,

together with the electrical conductivity.

Figure 7. Reactor Setup: Schematic overview of the lab-scale membrane bioreactor.


26

Figure 8. Reactor Setup: Lab-scale membrane bioreactors.

2.2. INOCULUM

The two membrane bioreactors were inoculated with activated sludge from the B-stage of the waste

water treatment plant Nieuwveer, in Breda (The Netherlands). MBR 1 and 2 were both inoculated

with 5.7 g VSS L-1.

2.3. OPERATION

MBR 1 and 2 were operated with a constant influent flow rate of 0.500 L d-1 and increasing nitrogen

concentrations (from 10-100% hydrolyzed urine). Influent was dosed every hour to spread the

nitrogen influent load in time. MBR 1 was operated at a constant salt concentration of a 100%

hydrolyzed urine solution (60-70 mS cm-1). MBR 2 was operated without salinity adaptation and the

initial EC was 5 mS cm-1 which is the normal working conductivity of the B-sludge from Breda.

Effluent was checked every morning and evening using indicator strips for total ammonia nitrogen

(TAN) and nitrite. If TAN (< 12 mg N L-1) or NO2- (< 6 mg N L-1) was low, the influent concentration was

increased. If TAN (> 12 mg N L-1) or NO2- (> 6 mg N L-1) was high, the influent concentration was

decreased or turned off.

Table 9 gives an overview of the initial and target operation parameters of the 2 MBR’s.


27

Table 9. Overview of the initial and target operation parameters (Start-up*/Target values**).

Operation parameter MBR 1 MBR 2

Working volume (L-1) 8.0 8.0 Q gas flow (L min-1) 5 5

EC (mS cm-1) 70 5*/70**

pH 6.9 – 7.1 6.9 – 7.1

Temperature (°C) 20 20

Nitrogen loading rate (gN L-1 d-1) 0.05*/0.5** 0.05*/0.5**

Influent concentration (gN L-1) 0.797*/7.97** 0.797*/7.97**

HRT (d) 16 16

2.4. SYNTHETIC HYDROLYZED URINE

The reactor influent was based on standard concentrations for urine (Udert et al., 2003b). Sodium

acetate was added to serve as a COD source (COD concentration: 8 g L-1 or 0.125 M acetate). The

composition of the synthetic urine after ureolysis with organic substances is given in Table 10.

Table 10. Composition of synthetic hydrolyzed urine.

Parameter Quantity (g L-1) Concentration (M)

NH3 tot 9.674 0,568

PO4 tot 1.662 0,0175

Acetate 7.506 0,125 CO3 tot 16.26 0,271

Cl 4.183 0,118 SO4 1.556 0,0162

Na 2.552 0,111

K 2.201 0,0563 Alkalinity 0,584 Ionic strength 0,637 pH [-] 8,87

3. BATCH EXPERIMENTS

3.1. EFFECT OF SALT CONCENTRATION ON NITRIFICATION ACTIVITY

To investigate the effect of salt concentration on the nitrification activity of the AOB’s and NOB’s in

the HANDS sludge, a batch experiment in Erlenmeyer flasks (250 ml) was performed. For every salt

concentration (1, 2, 3.5 and 5 g L-1 NaCl) tests were performed in triplicate.

In every Erlenmeyer 75 ml of medium was added together with 25 ml of washed and activated

HANDS sludge. Activation of the nitrifiers was managed by dosing 50 mg L-1 NH4+-N, 50 mg L-1 NO2

--N,

15 mg L-1 PO4-P and incubation on 20°C for 24 hours on a shaker (111 rpm). The biomass was washed


28

3 times for 5 minutes on 10.000 rpm with distilled water that was brought on the same conductivity

as the supernatant after washing (physiological conditions). The medium consisted out of the

medium used for growing the different algae species and was brought to pH 6 with a phosphate

buffer. The medium contained 50 mg NO2--N L-1 in the form of sodium nitrite and 50 mg NH4

+-N L-1 in

the form of ammonium chloride.

3.2. EFFECT OF AMMONIUM AND SALT CONCENTRATION ON ALGAE GROWTH

To investigate the influence of the salt and ammonium concentration on the growth of different

microalgae species a batch experiment was performed. All species and a mix of all species together

were exposed to 8 different media and tests were performed in quadruplicate. The experiment was

realized in four 96 well micro titer plates of which the first 2 where used to examine the effect of 4

salt concentrations (1, 2, 3.5 and 5 g L-1 NaCl) and the second 2 to examine the effect of 4 ammonium

concentrations (50, 100, 200 and 1000 mg NH4+-N L-1 in the form of ammonium nitrate). The media

consisted out of the normal microalgae culture media brought to a pH of 6 with a phosphate buffer.

The division of the 96 well micro titer plates is given in Figure 9. All species were given a number

from 1 to 9 and B stands for Blanco. The blanks were spread over the plate to eliminate influences of

place specificity.

Figure 9. Division of the 96 well micro titer plates.

B B B B 7 7 7 7 B B B B

1 1 1 1 8 8 8 8 4 4 4 4

2 2 2 2 9 9 9 9 5 5 5 5

3 3 3 3 B B B B 6 6 6 6

4 4 4 4 B B B B 7 7 7 7

5 5 5 5 1 1 1 1 8 8 8 8

6 6 6 6 2 2 2 2 9 9 9 9

B B B B 3 3 3 3 B B B B

Each well had a volume of 300 µl and after filling the wells with growth media, each well was

inoculated with 10–30 μL of microalgae culture to obtain an optical density at 620 nm (OD 620) of

0.05 – 0.10. The plates were incubated in an orbital shaking incubator at 20°C for 160 hours. The

optical density was measured every day with a plate reader (spectrophotometer platereader Tecan

Infinite) at 620 nm (to measure chlorophyll a and chlorophyll b which are the main chlorophyll

pigments in Chlorella). Before measurement, the micro titer plates were shook during 1 minute at

the linear amplitude of 2 mm.

3.3. NITRIFICATION ACTIVITY TEST PROTOCOL

To standardize the nitrification activity tests, a protocol was optimized. The working volume was 100

mL which was brought into 250 mL Erlenmeyer flasks, leaving a headspace volume equal to or

greater than the volume of the working volume, to ensure aerobic conditions. The working volume


29

existed out of 5 mL of synthetic substrate medium, a phosphate buffer solution that comprised

NaHCO3 to provide enough alkalinity, the used biomass and distilled water. Stoichiometric there is a

need for 12 g NaHCO3 per g N nitrified. The Erlenmeyer flasks were placed at a shaker at 20°C in a

temperature controlled room and 4 mL samples were taken every hour, filtered over a 0.45 µm

syringe filter and stored in the fridge at 4 °C. The pH and DO were measured every hour and the

samples were analyzed for nitrite, nitrate and ammonium. Biomass was prepared by centrifuging the

biomass and washing once with tap water. Every Erlenmeyer was inoculated with 5 g of centrifuged

biomass. The initial and final TSS and VSS concentrations were determined on 5 g of biomass and 10

mL working volume respectively. Also the initial and final electrical conductivity was measured.

4. ANALYTICAL TECHNIQUES

4.1. AMMONIUM

4.1.1. TESTSTRIP

An indication of the ammonium concentration was found with teststrips for ammonium (MQuant

ammonium test). These strips have a measuring range between 10 en 400 mg NH4+

L-1 or 7.8 – 310

mg NH4+-N L-1. 5 mL of sample is filtered over a 0.45 µm Chromafil Xtra filter (Machery-Nagle, PA,

USA) after which 10 drops of NaOH is added.

4.1.2. NESSLER COLORIMETRIC METHOD

Daily samples of the influent and the bulk of the reactor were filtered over a 0.45 µm Chrimafil Xtra

filter (Machery-Nagle, PA, USA) and placed in the dark by 4°C. The concentration of ammonium was

determined spectrophotometric following the standard method of Nessler (Greenberg et al. 1992). In

a first step, possible interfering elements like manganese and iron are captured by KNa-tartraat

(KNaC4H4O6.4H2O). In a second step the Nessler reagens is added. This is an alkaline solution with

HgI42-ions, which forms a yellow collared complex with ammonium. The intensity of the yellow color

is directly linked with the concentration of ammonium in the sample. The used spectrophotometer

was a Biochrom WPA Lightwave II (Biochrom Ltd, Cambridge, UK). Measurements were done by a

wavelength of 425 nm and a measuring range of 0 – 5 mg NH4+

L-1. Some organic aliphatic or aromatic

compounds such as amines, chloramines, hydrazines and aldehydes can cause interferences.

4.1.3. STEAM DISTILLATION

Total ammonium nitrogen (TAN) was analysed by steam distillation according to Standard methods

(4500-NH3 B; APHA, 1992). The sample was diluted in a distillation tube depending on the suspected

ammonium concentration to remain between the detection limits (5 – 300 mg NH4+-N L-1). 0.4 mg

MgO was added to create weak alkaline conditions and the ammonia was captured in a boric acid

indicator (as (NH4)3BO3) after distillation. Afterwards the captured ammonia is titrimetrically

determined with hydrochloric acid 0.02 M. Titration was carried out with a pH meter.


30

4.2. TOTAL KJELDAHL NITROGEN

Kjeldahl nitrogen was analysed according to Standard methods (4500-Norg B; APHA, 1992). Organic

nitrogen was determined as the difference between Kjeldahl nitrogen and ammonium nitrogen. The

sample was diluted in a Kjeldahl tube so that the concentration was located between the detection

limits (9 - 250 mg NH4+-N/L). The organic nitrogen present in the sample is transformed into

ammonium nitrogen (NH4)2SO4 by means of destruction at 400°C for 1.5 hours with sulphuric acid

(H2SO4) (98%) and potassium and copper sulphate (K2SO4, CuSO4) as a catalyst. The digested sample is

distillated and ammonia is captured in a boric acid indicator. The ammonia which is captured in that

acid solution (as (NH4)3BO3) is titrimetrically determined with hydrochloric acid 0.02 M.

4.3. NITRITE AND NITRATE

4.3.1. TESTSTRIP

An indication of the nitrite and nitrate concentration was found with teststrips (MQuant nitrite and

nitrate test). These strips for nitrite and nitrate have a measuring range between 0 and 80 mg NO2- L-1

or 0 – 24 mg NO2--N L-1. 5 mL of sample is filtered over a 0.45 µm Chromafil Xtra filter (Machery-

Nagle, PA, USA).

4.3.2. ION CHROMATOGRAPHY

For determination of nitrite and nitrate concentrations, the samples were diluted tenfold with milliQ

and analysed with an ionchromatograph (IC 761, Compact, Methrom AG, Swiss) equipped with a

metrosep A supp 5 column and a conductivity meter. NaCO3 (1.06 g L-1) was used as eluent with a

flow rate of 0.7 mL min-1 and a sample loop of 20 µL. The surface under the peak is proportional with

the concentration in the sample. The detection limit of this equipment ranges from 0.05 – 100 mg L-1

nitrite or nitrate.

4.4. CHEMICAL OXYGEN DEMAND

The chemical oxygen demand was determined photometrical with van Nanocolor® COD 160 testkits

(Machery-Nagel, PA, USA). The COD concentration was determined by a 2 hours lasting oxidation

with potassium dichromate/sulfuric acid, catalyzed by silver and by a temperature of 148 °C. For this

analysis, 2 mL of sample was filtered over a 0.45 µm Chromafil filter and added to the solution. The

Nanocolor® Vario 4 was used for destruction of the samples. After cooling down to room

temperature, the COD was determined with a digital photometer Nanocolor® 500 D. The COD 160

testkits permit to determine the oxygen demand within a range of 15 to 160 mg O2 L-1.

4.5. PH

The pH was measured continuously using a digital pH measuring device (Dulcometer with pH

electrode Dulcotest PHEP-112, Prominent). Calibration was done with a standard buffer solution with

pH 4 and pH 7.


31

4.6. DISSOLVED OXYGEN

Dissolved oxygen concentration was measured on a daily base with a portable, digital oxygen probe

(Hach HQ40d portable meter). The concentration was expressed in mg O2 L-1.

4.7. ELECTRICAL CONDUCTIVITY

Electrical conductivity was measured on a weekly base with a bench-top conductivity analyzer with

EC-electrode (Consort C833 multi-channel analyzer). The measurements were automatically

temperature corrected for 20°C.

4.8. TOTAL SUSPENDED SOLIDS AND VOLATILE SUSPENDED SOLIDS

The biomass concentration was determined weekly based on the amount of the volatile suspended

solids (VSS). This concentration was found by substracting the inorganic solids from the total

suspended solids (TSS). TSS and VSS were performed according Standard Methods 2540D and E

(APHA, 1997). Samples were taken from the bulk of the reactor. 10 mL of the sample is brought on a

glass-fiber filter which is mounted onto a vacuum filtration system. The filter is dried for at least 1

hour in an oven at 105°C. What is left on the filter is the TSS. The residue from the TSS method is

ignited to constant weight at 550°C for 1.5 hours. The residue on the filter is the VSS.

To determine the initial TSS and VSS concentration for the activity tests, solids analyses were

performed by centrifugation of a known sample volume and by weigh difference of the pellet after

drying at 105°C (TSS) and incinerating at 600°C (VSS).

4.9. HEADSPACE: OXYGEN, NITROGEN AND CARBON DIOXIDE

The gas phase composition of the photobioreactor headspace was analyzed with a Compact GC

(Global Analyser Solutions, Breda, The Netherlands), equipped with a Molsieve 5A pre-column and

Porabond column (CH4, O2, H2 and N2) and a Rt-Q-bond pre-column and column (CO2, N2O and H2S).

Concentrations of gases were determined by means of a thermal conductivity detector.

5. MICROSCOPY

Samples from the different reactors were analyzed with light microscopy. A Zeiss microscope (type

Axioskop 2 plus) was used.

PART 3 Results

32

PART 3: RESULTS

1. PHOTOBIOREACTOR (PBR)

The goal of this photobioreactor was to oxidize ammonium to nitrate in synthetic urine. To do so, a

consortium of microalgae and nitrifying bacteria was developed, in which the microalgae produced

the oxygen in-situ for the nitrification process.

1.1. PRELIMINARY BATCH TESTS

1.1.1. EFFECT OF SALT CONCENTRATION ON NITRIFICATION ACTIVITY

As urine has a high salt concentration (an electrical conductivity of 18 mS cm-1), the effect of different

salt concentrations was investigated on the nitrification activity of the HANDS (highly active

nitrification denitrification sludge) inoculum and on microalgal growth.

Figure 10 shows the effect of salinity on the nitrification activity of AOB and NOB in the HANDS

sludge. Batch experiments were performed with four different salinities: 1 g L-1 NaCl (4.54 mS cm-1), 2

g L-1 NaCl (6.55 mS cm-1), 3.5 g L-1 NaCl (9.95 mS cm-1) and 5 g L-1 NaCl (13.60 mS cm-1). Tests were

performed in triplicates.

The AOB did not show any nitrification activity during the experiment. It is not clear if this was due to

inactive inoculum or caused by the low pH. The Erlenmeyer flasks were buffered at a pH of 6

however due to insufficient buffer capacity, the pH rose to a final value of 6.8 (on average). In

contrast to the AOB, the NOB showed an increasing activity with increasing salinity. NOB activities

were calculated from the first three data points and are given in Table 11. Dissolved oxygen was kept

higher than 6 mg O2 L-1 which means that no anoxic conditions were present which was confirmed by

the nitrogen mass balance. The NOB activity was used to calculate the necessary reactor VSS

concentration to nitrify the total target nitrogen load (see further).

Table 11. Nitrification activities of HANDS sludge for different electrical conductivities.

Salt addition

(g NaCl L-1)

pH* EC

(mS cm-1)

AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3-- N gVSS-1 d-1)

1 6.68 ± 0.09 4.54 0 8.57 ± 0.10 2 6.67 ± 0.13 6.55 0 9.84 ± 1.38

3.5 6.65 ± 0.13 9.95 0 9.91 ± 1.28

5 6.63 ± 0.15 13.60 0 13.88 ± 0.26 (*) average pH values during the entire experiment.

PART 3 Results

33

Figure 10. Effect of different salt concentrations on HANDS activity. Top left: 1 g NaCl L-1. Top right: 2

g NaCl L-1. Bottom left: 3.5 g NaCl L-1. Bottom right: 5 g NaCl L-1.

1.1.2. EFFECT OF AMMONIUM AND SALT CONCENTRATION ON ALGAE GROWTH

In this experiment the influence of different salt and ammonium concentrations on the growth of

different microalgal species was determined. The optical density (OD) at 620 nm served as a

measurement for growth. The different species were grown separately and all together in 8 different

media. Tests were performed in quadruplicate.

Figure 11 displays the optical density at 620 nm for the algae mix in function of time for different salt

(left) and ammonium (right) concentrations. The results for all microalgal species separately are

given in the appendix.

0

10

20

30

40

50

0 20 40 60 80 100 120

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

10

20

30

40

50

0 20 40 60 80 100 120

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

10

20

30

40

50

0 20 40 60 80 100 120

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

10

20

30

40

50

0 20 40 60 80 100 120

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

34

Figure 11. OD at 620 nm in function of time for different salt and ammonium concentrations Left:

effect of salinity on the algal mix; Right: effect of ammonium concentration on the algal mix.

The microalgal mix (all microalgae species together) showed a higher growth speed than each of the

species separately. Optical densities between 0.65 and 1.25 were reached depending on ammonium

and salt concentration. No optical densities higher than 0.4 were observed for the single species.

Higher salt concentrations (≥ 3.5 g NaCl L-1) delayed the rapid growth phase of the algae mix.

Regarding the effect of ammonium concentration, the figure indicates that a lower ammonium

concentration of 50 mg NH4+-N L-1 was beneficial for algae growth in comparison with the higher

ammonium concentrations.

From this experiment, concerning algae growth, we conclude that if best possible algal growth rates

are required in the reactor, salt concentrations lower than 3.5 g NaCl L-1 (9.95 mS cm-1) and

ammonium concentrations lower than 50 mg NH4+-N L-1 are favorable.

1.2. PBR NITRIFICATION REACTOR

Prior to reactor inoculation the theoretical oxygen production and consumption of microalgae and

nitrifying sludge was calculated. The target for the final nitrogen loading rate was chosen to be 250

mg N L-1 d-1. This choice was based on a reported maximum observed nitrification rate of 185 mg

NH4+-N L-1 d-1 in a similar non-optimized system (Karya et al., 2013). Stoichiometrically, 4.57 g O2 is

needed to oxidize 1 g of nitrogen. Assuming a maximum photosynthetic oxygen production rate of

0.14 mg O2 mg-1 algal-TSS h-1 at 20°C (Drapcho and Brune, 2000), it was possible to calculate that the

algae TSS concentration must be at least 0.43 g TSS L-1. The necessary TSS concentration of nitrifying

HANDS sludge was calculated based on the NOB rate of 14 mg NO2--N gVSS-1 d-1 obtained from the

preliminary activity tests on HANDS sludge (Results; paragraph 1.1.1) and the final reactor

nitrification capacity of 250 mg N L-1 d-1. Based on this, the necessary TSS concentration of nitrifying

HANDS sludge was 17.9 g VSS L-1. The reactor was inoculated with the available biomass and initial

VSS concentration was 0.7 g VSS L-1.

Figure 12 gives an overview of the different nitrogen species of both influent and effluent, the

nitrogen loading rate (gN L-1 d-1) and the nitrification efficiency (%) during photobioreactor operation.

0

0,2

0,4

0,6

0,8

1

1,2

1,4

1,6

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

0

0,2

0,4

0,6

0,8

1

1,2

1,4

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4+-N/L

100 mg NH4+-N/L

250 mg NH4+-N/L

1000 mg NH4+-N/L

PART 3 Results

35

Figure 12. Nitrogen species in influent and effluent, nitrogen loading rate (gN L-1 d-1) and nitrification efficiency (%) during PBR operation.

0

10

20

30

40

50

60

70

80

90

100

0

100

200

300

400

500

600

700

0 10 20 30 40 50 60 70 80 90 100 110 120 130 140 150 160 170

Nit

rifi

cati

on

eff

icie

ncy

(%

); N

itro

gen

load

ing

rate

(m

g N

L-1

d-1

)

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (d)

NH4+ effluent NO2- effluent

NO3- effluent Organic nitrogen effluent

Kjeldahl nitrogen influent Nitrification efficiency (%)

Nitrogen loading rate (gN L d)

I II

III

IV

V VI

PART 3 Results

36

Figure 13 visualizes the influent and effluent COD concentration together with the COD removal

efficiency during PBR operation. Figure 14 visualizes the dissolved oxygen (DO), TSS and VSS

concentrations in function of time during operation.

Figure 13. Influent and effluent COD concentrations and COD removal efficiency during PBR

operation.

Figure 14. Dissolved oxygen, TSS and VSS concentration in function of time during PBR operation.

Although the reactor achieved nitrogen removal efficiencies between 60 and 100% (total Kjeldahl

nitrogen in the influent compared to the effluent concentration), the first 55 days, no net nitrification

(no nitrate in the effluent) was observed. During the first 50 days of reactor operation, three changes

were made. At day 15, a Kjeldahl nitrogen concentration of 246 mg N L-1 (89 mg NH4+-N L-1 and 157

0

10

20

30

40

50

60

70

80

90

100

0

100

200

300

400

500

600

700

800

900

0 20 40 60 80 100 120 140 160

CO

D r

emo

val e

ffic

ien

cy (

%)

Co

nce

ntr

atio

n (

mg

O2

L-1)

Time (days)

COD influent

COD effluent

COD removal efficiency (%)

0

1

2

3

4

5

6

7

8

9

0 20 40 60 80 100 120 140 160

mg

O2

L-1; g

TSS

L-1;

gVSS

L-1

Time (days)

Dissolved oxygen (mg O2/L)

gTSS/L

gVSS/L

PART 3 Results

37

mg Org-N L-1) accumulated and the reactor was reinoculated with 0.5 L ABIL (Ammonium Binding

Inoculum Liquid; Avecom, Belgium) and 0.5 L HANDS (highly active nitrification denitrification sludge;

Avecom, Belgium) to stimulate nitrification (Figure 12; I). Although a low nitrate concentration was

observed in the effluent (11 mg NO3--N L-1) and nitrification was still not achieved, the ABIL sludge

gave good settling properties to the biomass. VSS/TSS ratio decreased from 0.83 to 0.42 gVSS gTSS-1

after the addition of ABIL.

To investigate whether the biomass was inactive or the reactor medium had an inhibitory effect,

nitrification activity batch tests were performed at day 19 of reactor operation. Centrifuged biomass

was supplied with fresh medium and activated HANDS sludge was supplied with the biomass

supernatants. Also some reactor biomass without adaptations was placed on the shaker to see if

reactor conditions or operation were a limiting factor. The test was performed without repeats.

Figure 15 shows the nitrification results for the 3 configurations.

Figure 15. Ammonium, nitrite and nitrate profiles. Upper left: HANDS + reactor supernatans; Upper

right: PBR biomass + fresh medium; Bottom: PBR biomass + reactor supernatans.

For the three configurations, AOB and NOB activities were calculated based on ammonium removal

and nitrate formation respectively. In Table 12 the nitrification activities are listed together with the

0

20

40

60

80

100

0 50 100 150 200

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

20

40

60

0 50 100 150 200

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

20

40

60

80

100

120

0 50 100 150 200

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

38

initial VSS concentrations and electrical conductivity. DO concentrations were always higher than 5

mg O2 L-1.

Table 12. Nitrification activities for the three different configurations. HANDS + reactor supernatant;

PBR biomass + fresh medium; PBR biomass + PBR supernatant.

EC

(mS cm-1)

pH VSS

(g L-1)

AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

HANDS + reactor supernatant 9.45 7,34 ± 0,45 1.35 27.9 12.0

PBR biomass + fresh medium 8.92 7,23 ± 0,51 1.71 2.22 2.56

PBR biomass + PBR supernatant

8.69 7,06 ± 0,63 2.46 7.33 2.24

As can be seen from these results, the PBR biomass had a nitrification potential of 7.33 mg NH4+-N

gVSS-1 d-1 and 2.24 mg NO3--N gVSS-1 d-1. Good nitrification activity was observed in the nitrification

test with the HANDS and consequently the nitrification activity was not inhibited by an internal

reactor component. During this experiment, the dissolved oxygen in the erlenmeyer flasks did not

drop below 5 mg O2 L-1, while the reactor oxygen concentration was below 1.4 mg O2 L-1 after

influent dossage.

At day 25, aeration was started at a flow rate of 5 L h-1 compressed air (0.25 L O2 Lreactor-1 h-1) (Figure

12; II). This resulted in Kjeldahl nitrogen effluent concentrations of 0 mg N L-1 at day 34, but no

nitrification was initiated. At this moment the nitrogen loading rate was increased from 50 mg N L-1 d-

1 to 70 mg N L-1 d-1 (Figure 12; III). During this startup period, COD removal efficiencies fluctuated

between 61 and 93% while the COD loading rate was fluctuating between 53 and 93 mg O2 L-1 d-1

(Figure 13). The COD removal efficiency did not increase after compressed air addition. Biomass VSS

concentration increased from 0.33 to 3.03 g VSS L-1 during the first 26 days, after compressed air was

applied it decreased to 1.86 g VSS L-1 at day 40.

After 60 days of reactor operation, nitrification initiated without any changes in reactor operation.

Starting from day 110, the reactor reached nitrification efficiencies higher than 50% and at day 118,

compressed air was shut off (Figure 12; IV). The microalgal oxygen production could provide enough

oxygen to nitrify on complete photo-aeration (nitrification efficiencies decreased from 69 to 52% but

increased again to 64% after 7 days). Oxygen levels were highly variable but remained above 6 mg O2

L-1. When nitrification initiated, COD removal efficiencies dropped slightly towards values between 44

and 83%.

At day 125 the nitrogen loading rate was increased from 70 mg N L-1 d-1 to 90 mg N L-1 d-1 by reducing

the influent dilution to 12% synthetic urine (Figure 12; V). Finally at day 134 the nitrogen loading

rate was increased from about 90 mg N L-1 d-1 to about 100 mg N L-1 d-1 (15% dilution of synthetic

urine) until day 136 (Figure 12; VI). After that the reactor nitrification performance dropped sharply

PART 3 Results

39

with nitrification efficiencies lower than 30% and VSS concentration decreases from 2.68 to 1.81

from day 133 to day 158 (Figure 14). Also low dissolved oxygen concentrations were detected.

VSS/TSS ratio remained constant around 0.84 gVSS gTSS-1.

During the period from day 90 until day 134, the reactor nitrification efficiency was located between

38 and 69% and although a settling time of 10 minutes was foreseen, no visual settling took place.

The remaining nitrogen was fixed by the algae and washed out. Calculations showed that, without

optimization of the settling, 21 – 27% of influent Kjeldahl nitrogen was removed with the algae

biomass (biomass washout) while this dropped to 13% when settling was optimized. If no nitrification

occurred, 20 – 52% of the influent Kjeldahl nitrogen disappeared with the effluent in the form of

ammonium and organic nitrogen (first 55 days) and in the form of nitrate when nitrification occurred.

For example, on day 103, the nitrification efficiency was 50%, 21% of the influent nitrogen

concentration disappeared with the effluent in the form of biomass and 13% accumulated in the

reactor in the form of nitrate. This calculation did not account for 16% of the nitrogen influent

concentration.

During day 133, the sludge volume index (SVI) was determined. SVI5 was 90.32 ml g-1 and SVI30 was

58.06 ml g-1 which indicates a sufficiently well settling capacity.

A second nitrification activity test was performed at day 146, after the decrease in nitrification

efficiency. The goal of this test was to determine whether the biomass was nitrifying at full capacity,

or whether oxygen or carbon limitations in the reactor were limiting nitrification activity. Figure 16

shows the nitrification results.

Figure 16. Activity test on the PBR biomass after 146 days.

AOB and NOB nitrification activity was calculated based on ammonium consumption and nitrate

production respectively. The results are presented in Table 13. Dissolved oxygen was always higher

than 2 mg O2 L-1.

0

5

10

15

20

25

30

35

40

45

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4NO2NO3

PART 3 Results

40

As the VSS concentration in the reactor was 1.81 g VSS L-1, the AOB in the reactor have a nitrification

potential of 121.9 mg NH4-N L-1 d-1 while the NOB can produce 86.61 mg NO3-N L-1 d-1.

Between day 146 and 167, nitrification efficiency fluctuated between 0 and 30%. At day 168, the

nitrogen loading rate was decreased from 90 mg N L-1 d-1 to 35 mg N L-1 d-1 (5% dilution of synthetic

urine) and nitrification efficiency increased to 77%. From day 168 until 180, nitrification efficiency

remained between 52% and 53%. After day 146, the COD removal efficiencies dropped slightly and

fluctuated between 20 and 67%.

After 180 days of reactor operation, a last activity test was performed on the PBR biomass. The goal

was to investigate the nitrification potential of the PBR biomass and compare it with the nitrification

activity in the bioreactor. The test was set up in the same way as the previous nitrification activity

tests. Figure 17 shows the results.

Figure 17. Activity test on the PBR biomass after 180 days.

AOB and NOB nitrification activity was calculated based on ammonium consumption and nitrate

production respectively. Dissolved oxygen was always higher than 7 mg O2 L-1. The results are

presented in Table 13.

Table 13. PBR biomass nitrification activities after 146 days and after 180 days, pH during the

experiment and initial VSS concentration.

Sampling

time (day)

pH VSS

(g L-1)

AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

146 7.02 ± 0.11 2.22 ± 0.55 67.35 ± 5.63 47.85 ± 0.82 180 7.06 ± 0.02 3.93 ± 0.56 11.46 ± 0.14 10.54 ± 2.88

0

5

10

15

20

25

30

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

41

1.2.1. DISSOLVED OXYGEN PROFILE OVER ONE CYCLE

At day 133, dissolved oxygen (DO) levels were monitored during 1 cycle of 8 hours (Figure 18). In the

beginning of the cycle, the DO concentration was 7.41 mg O2 L-1. Lights were turned off and the

reactor was wrapped in tin foil to observe the oxygen consumption in the reactor. When the DO

reached values of less than 3 mg O2 L-1, lights were turned on again. During the following period, the

concentration of oxygen remained the same but nitrification was still ongoing (ammonium and nitrite

concentrations did not reach zero yet), which indicates a simultaneous oxygen production and

consumption. From the moment that ammonium and nitrite were fully oxidized (reactor

concentration was monitored with ammonium and nitrite test strips), the oxygen concentration

started to increase again.

Figure 18. Dissolved oxygen profile during 1 cycle.

1.3. MICROSCOPY ON PBR BIOMASS

Figure 19 presents the light microscopy images of the PBR biomass after 180 days of reactor

operation. Based on these pictures it is possible to distinguish 4 species: Scenedesmus sp. (green

algae), Chlorella sp. (green algae), Synechocystis sp. (cyanobacteria) and Leptolyngbya sp.

(cyanobacteria). Between the large microalgae cells, the bacteria are present in low numbers. We

can say that a selection took place towards one dominant algae species (Scenedesmus sp.).

2

3

4

5

6

7

8

0 1 2 3 4 5 6 7

Dis

solv

ed o

xyge

n (

mg

O2

L-1)

Time (h)

Light off Light on

Ammonium and nitrite consumed

PART 3 Results

42

Figure 19. Microscopy images of PBR biomass. Top: 400 x magnification; Bottom: 1000 x

magnification. (1) Chlorella sp.; (2) Scenedesmus sp.; (3) Synechocystis sp.; (4) Leptolyngbya sp.;

(5) Bacteria.

1

1

2

3

4

5

PART 3 Results

43

2. MEMBRANE BIOREACTORS (MBR)

The goal of this study was to determine the most optimal start-up strategy for urine nitrification

membrane bioreactors.

2.1. NITRIFICATION ACTIVITY TEST B-SLUDGE

Before the two membrane bioreactors were inoculated, the nitrification activity of the activated B-

sludge, which was the inoculum, was determined. This activated B-sludge originated from the B-

stage of the waste water treatment plant Nieuwveer, in Breda (the Netherlands). The nitrification

activity of the biomass was then used to determine the necessary VSS concentration in the reactor to

nitrify all nitrogen of the target nitrogen loading rate of 500 mg N L-1 d-1.

This test was performed following the optimized nitrification activity test protocol described earlier.

Two different substrate media were used. One contained 50 mg NO2--N L-1, the other one contained

50 mg NH4+-N L-1 and 50 mg NO2

--N L-1. The whole was brought on pH 7 with a phosphate buffer and

inoculated with 5 g of PBR biomass. Tests were performed in triplicates. Figure 20 shows the results.

Figure 20. Nitrogen species profiles of the preliminary activity test on inoculum B-sludge.

NOB activity was calculated based on the first and second graph (nitrate production). AOB activity

was calculated based on the results from the second graph (NH4+ removal). The nitrification activity is

presented in Table 14. The average pH value was 7.58 despite the addition of the phosphate buffer

that aimed at a pH of 7. Average dissolved oxygen concentrations were 2.19 and 1.64 mg O2 L-1.

Table 14. Nitrification activity B-sludge, pH and DO during the experiment and initial VSS

concentration.

Initial nitrogen

concentration

pH DO

(mg O2 L-1)

VSS

(g L-1)

AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

50 mg NH4+-N L-1 7.58 ± 0.23 2.19 ± 0.94 3.66 (n/a) 50.5 ± 2.08

50 mg NH4+-N L-1

50 mg NO2--N L-1

7.55 ± 0.29 1.64 ± 0.22 3.66 21.3 ± 0.91 59.9 ± 3.49

0

10

20

30

40

50

0 2 4 6

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4

NO2

NO3

0

10

20

30

40

50

0 2 4 6

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4

NO2

NO3

PART 3 Results

44

Based on these results it can be seen that ammonium oxidation is the rate limiting step in this

activated sludge nitrification process. When a target of 500 mg N L-1 d-1 is the pursued target

nitrification rate, a reactor biomass concentration of 23.51 g VSS-1 L-1 is necessary based on this AOB

activity or 7.25 g VSS L-1 based on the NOB activity. The two membrane bioreactors were inoculated

with the complete volume of available biomass which resulted in a VSS concentration of

approximately 5.70 g VSS L-1.

2.2. MBR NITRIFICATION REACTORS

2.2.1. MBR 1: HIGH SALINITY START-UP STRATEGY

Figure 21 visualizes the nitrogen species in membrane bioreactor 1 in function of time while Figure

22 visualizes the free ammonia (FA) concentration. Free nitrous acid concentrations remained below

0.020 mg N L-1 and did not form any threat towards AOB nor NOB. Due to the conductivity shock of

70 mS cm-1, the biomass lost total activity during the first 12 days. Ammonium accumulated to a

concentration of 126 mg NH4+-N L-1 and free ammonia to a concentration of 0.65 mg N L-1 which is an

inhibitory concentration for the NOB. At day 12, nitrite accumulation in the reactor indicated AOB

activity. From day 23 on, nitrate accumulation occurred while nitrite concentrations decreased,

indicating increased NOB activity. At that moment, FA concentrations decreased to 0.08 mg N L-1.

Figure 21. Nitrogen species in the effluent for MBR 1.

0

50

100

150

200

250

300

350

0

20

40

60

80

100

120

140

0 10 20 30 40 50 60

mg

NO

3- -

N L

-1

Co

nce

ntr

atio

n (m

g N

L-1

)

Time (days)

NH4+NO2-NO3-

PART 3 Results

45

Figure 22. Reactor free ammonia (FA) concentrations for MBR 1.

Figure 23 represents the increasing nitrogen loading rate together with the electric conductivity. At

day 35, ammonium and nitrite reactor concentrations were practically zero and influent (10%

dilution of urine) was dosed again at a nitrogen loading rate of 0.05 gN L-1 d-1. From that moment

onwards, it was possible to increase the loading rate from 0 to 0.19 gN L-1 d-1 in 19 days by lowering

the dilution of the synthetic urine while maintaining the influent flow rate.

Figure 23. Nitrogen loading rate, influent urine dilution and EC in reactor effluent for MBR 1.

Figure 24 (Left) shows the TSS and VSS concentrations and the VSS/TSS ratio in MBR 1. During the

first 35 days, the VSS concentration decreased from 5.64 g VSS L-1 to 0.92 VSS L-1 after which it

increased again to 2.2 g VSS L-1 at the same moment influent was dosed again. The VSS/TSS ratio

folows the same trend as the VSS concentration; it reached its lowest point at 35 days (0.27 gVSS

gTSS-1) and then increased to a value of 0.48 gVSS gTSS-1. The initial VSS/TSS ratio of 0.5 gVSS gTSS-1

however, is much lower then the initial value in MBR 2 of 0.75 gVSS gTSS-1 (see further). This means

that a considerable part of VSS was submitted to imidiate lysis (due to the high salt shock). Figure 24

(Right) presents the nitrification efficiency which is negative at first, due to nitrate consumption that

originated from the biomass residual.

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0 10 20 30 40 50 60

con

cen

trat

ion

mg

N L

-1

Time (days)

Free ammonia

0

10

20

30

40

50

60

70

0

0,02

0,04

0,06

0,08

0,1

0,12

0,14

0,16

0,18

0,2

0 10 20 30 40 50 60

EC (

mS

cm-1

); U

rin

e d

iluti

on

(%

)

Nit

roge

n lo

adin

g ra

te (

gN L

-1 d

-1)

Time (days)

Nitrogen loading rate

Electrical conductivity

Synthetic urine dilution (%)

PART 3 Results

46

Figure 24. Left: TSS and VSS concentrations and the VSS/TSS ratio. Right: Nitrification efficiency.

During the first 13 days, the VSS concentration decreased from 5.64 g VSS L-1 to 1.56 g VSS L-1. During

these 13 days, the ammonium concentration increased until 126 mg NH4+-N L-1 although only 49 mg

NH4+-N L-1 was dosed with the influent. The remaining ammonium orriginated from biomass decay

(77 mg NH4+-N L-1). From biomass analysis it was seen that the total Kjeldahl nitrogen content of the

biomass is 108.35 mg N gVSS-1. 4.08 gVSS L-1 is the equivalent of 442 mg N L-1 but only 77 mg NH4+-N

L-1 was released into the reactor. This means that only 77 mg N L-1 was converted to ammonium.

After 21 days of reactor operation a nitrification activity test was performed. This test was conducted

to examine the nitrification activity at different conductivities of the sludge at that moment and to

compare these activities with the initial nitrification activity of the activated B-sludge. At the moment

of this activity test MBR 1 was inactive due to the salt shock. The test was performed following the

optimized nitrification activity test protocol described earlier. The working volume was 100 mL which

was brought into 250 mL flasks. The substrate medium contained 25 mg NH4+-N L-1 and 25 mg NO2

--N

L-1.

The biomass was tested at 5 mS cm-1 (initial conductivity of the B-sludge) and 70 mS cm-1 (electrical

conductivity in MBR 1). Results are presented in Figure 25.

Figure 25. Results nitrification activity test after 21 days. Left: MBR I, 5 mS cm-1; Right: MBR I, 70 mS

cm-1.

0

0,1

0,2

0,3

0,4

0,5

0,6

0

2

4

6

8

10

12

0 10 20 30 40 50

VSS

/TSS

rat

io

g V

SS L

-1; g

TSS

L-1

Time (days)

TSS

VSS

VSS/TSS ratio

-10

10

30

50

70

90

0 10 20 30 40 50

Nit

rifi

cati

on

eff

icie

ncy

(%

)

Time (days)

Nitrification efficiency (%)

0

10

20

30

40

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

10

20

30

40

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

47

AOB and NOB activities were calculated based on ammonium removal and nitrate formation

respectively. Nitrification activities together with initial VSS concentrations and average pH values

are given in Table 15. Dissolved oxygen concentration remained above 4 mg O2 L-1.

Table 15. Nitrification activity MBR 1 different conductivities at day 21.

Conductivity pH VSS

(g VSS L-1)

AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

5 mS cm-1 7.09 ± 0.07 3.17 ± 0.08 1,95

0,60

70 mS cm-1 6.85 ± 0.12 3.17 ± 0.08 4,32

1,42

Original (5 mS cm-1) 7.55 ± 0.29 3.66 21.3 ± 0.91 59.9 ± 3.49

A low, almost unnoticeable activity was observed. This was an indication of the incomplete

adaptation towards the higher salt concentration.

After 55 days of reactor operation, a second nitrification activity test was performed to examine the

nitrification activity at different electric conductivities and compare it with the initial nitrification

activity of the B-sludge and the nitrification activity after 21 days. At the moment of this activity test

MBR 1 was adapting to a load of 35% hydrolyzed urine. The test was performed following the

optimized nitrification activity test protocol described earlier. The substrate medium contained 50

mg NH4+-N L-1 and 50 mg NO2

--N L-1.

The biomass was tested at 3 different conductivities, 5 mS cm-1 (initial electrical conductivity of the

activated B-sludge), 15 mS cm-1 (electrical conductivity at which MBR 2 started to experience

nitrification problems) and 45 mS cm-1 (the electrical conductivity of both MBRs at that moment).

Results are presented in Figure 26.

0

10

20

30

40

50

60

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

10

20

30

40

50

60

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

48

Figure 26. Results nitrification activity test. Top left: MBR I, 5 mS cm-1; Top right: MBR I, 15 mS cm-1;

Bottom: MBR I, 45 mS cm-1;


respectively. Nitrification activities together with initial VSS concentration and average pH values are

given in Table 16. Dissolved oxygen concentration remained above 3.82 mg O2 L-1.


Conductivity pH VSS

(g VSS L-1)

AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

5 mS cm-1 7.00 ± 0.03 3.17 ± 0.08 64.88 ± 6.85 88.80 ± 5.77 15 mS cm-1 7.00 ± 0.03 3.17 ± 0.08 77.93 ± 7.88 87.96 ± 6.64

45 mS cm-1 6.99 ± 0.03 3.17 ± 0.08 67.03 ± 6.55 43.97 ± 5.53

Original (5 mS cm-1) 7.55 ± 0.29 3.66 21.3 ± 0.91 59.9 ± 3.49

Figure 27 presents the trend of the AOB and NOB activity in function of the salinity. AOB had an

optimal activity at 15 mS cm-1 which indicates a shift in optimum from 5 to 15 mS cm-1. NOB

however, showed a remarkable lower activity at 45 mS cm-1, indicating a slower adaptation.

0

20

40

60

80

100

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

49

Figure 27. AOB and NOB nitrification activity trend at different salinities for MBR 1.

Figure 28 shows the light microscopy images of the biomass from MBR 1 after 60 days with a 400 x

and a 1000 x magnification. The left picture focuses on one sludge floc, while the right picture gives a

detailed view of the floc structure.


magnification.

The sludge flocs have an irregular form and almost do not contain filamentous bacteria. The floc has

a compact structure which indicates the good settlability. The compact structure can be caused by

the bubble aeration (STOWA, 1999).

1.1.1. MBR 2: INCREASING SALINITY START-UP STRATEGY

MBR 2 was operated without salinity adaptation and initial electrical conductivity was 5 mS cm-1

which is the conductivity of the B-sludge from Breda. The reactor was operated with a constant

influent flow rate of 0.500 L d-1 and increasing nitrogen concentrations from 10 - 100% hydrolyzed

urine.

30

40

50

60

70

80

90

100

0 10 20 30 40 50

Act

ivit

y (g

NH

4+-N

gV

SS-1

d-1

)

Electrical conductivity (mS cm-1)

AOB activity MBR 1

NOB activity MBR 2

PART 3 Results

50

Figure 29 visualizes the nitrogen species in membrane bioreactor 2 in function of time while Figure

30 visualizes the free ammonia (FA) concentrations during reactor operation. Free nitrous acid

concentrations remained below 0.015 mg N L-1 and did not form any threat towards AOB nor NOB.

Figure 29. Nitrogen species in the effluent for MBR 2.

Figure 30. Reactor free ammonia (FA) concentrations for MBR 2.

Figure 31 represents the increasing nitrogen loading rate together with an increasing electrical

conductivity. During the frist 6 operation days, a nitrogen loading rate of 0.32 g N L-1 d-1 was already

reached. In this timeframe, the electrical conductivity increased to 15 mS cm-1 and the nitrifying

bacteria lost their nitrification activity, first visible in a nitrite peak at day 7 (Figure 29). Free ammonia

reached values of 0.023 mg N L-1 after 7 days and were even higher from day 14 unitl day 21 (with a

maximum value of 0.52 mg N L-1). From day 20 onwards, it was possible to increase the nitrogen

loading rate again until after 43 days, the target value of 0.5 gN L-1 d-1 was reached. The electrical

conductivity remained at a stable value of on average 45 mS cm-1.

0

0,5

1

1,5

2

2,5

3

3,5

4

0

20

40

60

80

100

120

0 10 20 30 40 50 60

g N

O3

-N/L

Co

nce

ntr

atio

n (m

g N

L-1

)

Time (days)

NH4+

NO2-

NO3-

0

0,1

0,2

0,3

0,4

0,5

0,6

0 10 20 30 40 50

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (days)

Free Ammonia

PART 3 Results

51

Figure 31. Nitrogen loading rate, influent urine dilution and electric conductivity for MBR 2.

Figure 32 (Left) shows the TSS and VSS concentrations and the VSS/TSS ratio. During the first 13 days,

the VSS decreases from 5.74 to 3.02 gVSS L-1 after which it stays constant untill day 36. After this

period, the VSS concentration increases again to the final VSS concentration of 5.22 gVSS L-1. The

VSS/TSS ratio stays relatively constant at 0.6. The initial VSS/TSS ratio of 0.76 gVSS gTSS-1 however, is

higher then the initial value in MBR 1 of 0.5 gVSS gTSS-1. Figure 32 (Right) visualizes the nitrification

efficiency. Nitrification efficiency mostly stays above 80%. After 44 days, when the reactor reached

the target loading rate of 500 mg N L-1 d-1, the nitrification efficiency stays around 100%. This

indicated full nitrification without nitrogen loss.

Figure 32. Left: TSS and VSS concentrations and the VSS/TSS ratio. Right: Nitrification efficiencies.

Full nitrification was also proven by the amount of base consumption. Full nitrification of one mole of

NH4+-N to NO3

--N consumes 1.98 moles of alkalinity as bicarbonate. To fully nitrify the daily 4 grams

of nitrogen in the influent, 0.56 mol of alkalinity is needed. On daily basis, the base consumption was

0.3 L of 1M KOH which is the equivalent of 0.3 mol alkalinity. This means that 0.26 mol of alkalinity

0

20

40

60

80

100

0,0

0,1

0,2

0,3

0,4

0,5

0,6

0 10 20 30 40 50 60

EC (

mS

cm-1

); U

rin

e d

iluti

on

(%

)

Load

ing

rate

(g

N L

-1 d

-1)

Time (d)

Nitrogen loading rate

EC

Synthetic urine dilution (%)

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0,8

0,9

0

1

2

3

4

5

6

7

8

9

10

0 10 20 30 40 50

VSS

/TSS

rat

io

g (V

SS; T

SS)

/L

Time (days)

TSSVSSVSS/TSS ratio

0

20

40

60

80

100

0 10 20 30 40 50

Nit

rifi

cati

on

eff

icie

ncy

(%

)

Time (days)

Nitrification efficiency (%)

PART 3 Results

52

originates from the urine itself which is 46% of necessary alkalinity for full nitrification. This value

agrees with the value of 41% reported by Chen (2009).

After 21 days of reactor operation a nitrification activity test was performed. This test was conducted

to examine the nitrification activity at different conductivities of the sludge at that moment and to

compare these activities with the initial nitrification activity of the activated B-sludge. At the moment

of this activity MBR 2 was adapting to a load of 50% hydrolyzed urine. The test was performed

following the optimized nitrification activity test protocol described earlier. The working volume was

100 mL which was brought into 250 mL flasks. The substrate medium contained 25 mg NH4+-N L-1 and

25 mg NO2--N L-1. The biomass was tested at 5 mS cm-1 and at 27 mS cm-1 (electrical conductivity of

MBR 2 at that moment). Results are presented in Figure 33.

Figure 33. Results of the nitrification activity test after 21 days. Left: MBR II, 5 mS cm-1; Right: MBR II,

27 mS cm-1.



given in Table 17. Dissolved oxygen concentration remained above 2 mg O2 L-1.


Conductivity pH VSS AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

5 mS/cm 7.09 ± 0.08 3.91 ± 0.15 11.12

49.13

27 mS/cm 6.94 ± 0.07 3.91 ± 0.15 11.45

24.35

Original (5 mS cm-1) 7.55 ± 0.29 3.66 21.3 ± 0.91 59.9 ± 3.49

After 55 days of reactor operation a second nitrification activity test was performed. This test was

conducted to examine the nitrification activity at different conductivities of the sludge at that

moment and to compare these activities with the initial nitrification activity of the activated B-

sludge. At the moment of this activity test MBR 2 was running on 100% hydrolyzed urine. The test

was performed following the optimized nitrification activity test protocol described earlier. The

substrate medium contained 50 mg NH4+-N L-1 and 50 mg NO2

--N L-1.

0

20

40

60

80

100

120

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

20

40

60

80

100

120

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

53

The biomass was tested at 3 different conductivities, 5 mS cm-1 (initial electrical conductivity of the

activated B-sludge), 15 mS cm-1 (electrical conductivity at which MBR 2 started to face nitrification

problems) and 45 mS cm-1 (the electrical conductivity of both MBRs at that moment). Results are

presented in Figure 34.

Figure 34. Results of the nitrification activity test after 55 days. Top left: MBR II, 5 mS cm-1; Top right:

MBR II, 15 mS cm-1; Bottom: MBR II, 45 mS cm-1.



given in Table 18. Dissolved oxygen concentration remained above 2.16 mg O2 L-1.


Conductivity pH VSS AOB activity (mg

NH4+-N gVSS-1 d-1)

NOB activity (mg

NO3--N gVSS-1 d-1)

5 mS cm-1 7.07 ± 0.02 3.91 ± 0.15 36.24 ± 1.02 104 ± 5.28 15 mS cm-1 6.98 ± 0.03 3.91 ± 0.15 46.39 ± 3.78 97.98 ± 8.07

45 mS cm-1 6.94 ± 0.04 3.91 ± 0.15 37.95 ± 2.96 113.31 ± 5.96

Original (5 mS cm-1) 7.55 ± 0.29 3.66 21.3 ± 0.91 59.9 ± 3.49

0

20

40

60

80

100

120

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

20

40

60

80

100

120

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

0

20

40

60

80

100

120

0 2 4 6 8

Co

nce

ntr

atio

n (

mg

N L

-1)

Time (h)

NH4+

NO2-

NO3-

PART 3 Results

54

Figure 35 presents the trend of the AOB and NOB nitrification activity in function of the salinity. AOB

followed a similar trend as did the AOB of MBR 1 at that moment with a shifted optimum at 15 mS

cm-1.

Figure 35. AOB and NOB nitrification activity trend at different salinities for MBR 2.

Figure 36 shows the light microscopy images of the biomass from MBR 1 after 60 days with a 400 x

and a 1000 x magnification. The left picture focuses on one sludge floc, while the right picture gives a

detailed view of the floc structure.


magnification.

The sludge flocs have an irregular form are formed around filamentous bacteria. The floc has a

compact structure which indicates the good settlability. The compact structure can be caused by the

bubble aeration (STOWA, 1999).

As can be seen from the images, the sludge flocs contain a large number of cauliflower-like

structures. These formations have the typical morphology of ammonium oxidizing bacteria clusters.

30

50

70

90

110

0 10 20 30 40 50

Act

ivit

y (m

g N

H4

+ -N

gV

SS-1

d-1

)

Electrical conductivity (mS cm-1)

AOB activity MBR 2

NOB activity MBR 2

PART 4 Discussion

55

PART 4: DISCUSSION

1. PHOTOBIOREACTOR

1.1. NITRIFICATION EFFICIENCY

During the reactor start-up period, different strategies were pursued to obtain a reactor with well

settling flocs and stable nitrification. During the first 55 days, it was not possible to trigger AOB and

NOB activity (no nitrite nor nitrate was produced), although nitrogen removal efficiencies between

60 and 100% were obtained. These high removal efficiencies without nitrification were possible due

to the low settling of the reactor biomass and by consequence biomass washout. The low settling

properties of the biomass made the short initial settling time of 10 minutes insufficient and not only

the slowest settling microorganisms were washed out, but all biomass. Sludge retention time (SRT)

was equal to the hydraulic retention time (HRT) of 6.7 days and the effluent VSS concentration was

equal to the reactor VSS concentration which fluctuated between 1.42 and 3.03 g VSS L-1 during the

first 55 days. The fast washout conditions made it impossible for the nitrifiers to maintain proper

biomass concentrations. Maximum specific growth rates reported at 20°C for AOB and NOB are

1.05 d−1 < μmax,AOB < 1.4 d−1 and 0.91 d−1 < μmax,NOB < 1.31 d−1 respectively (Munz et al., 2011). The

photobioreactor that was meant to stabilize nitrogen in source separated urine for further transport

and nutrient recovery became an efficient production unit for algal biomass.

It is possible that nitrification occurred but that all small amounts of nitrite and nitrate were directly

fixed by the algae. Objection could be that if acidifying nitrification occurred, alkalinity consumption

should appear. However, because carbon dioxide dosage (at 0.72 L h-1) may counteract the

acidification of the nitrification mechanism, almost no acid or base was consumed. Li et al. (2008)

reported that nitrate was the most favorable nitrogen source for the cell growth and lipid production

of Neochloris oleoabundans (green algae) among nitrate, urea and ammonium. Wu et al. (2013)

confirmed this and reported that if ammonium and nitrate are the present nitrogen sources,

microalgal cells would uptake ammonium first, and then uptake nitrate sequentially. It has been

researched that a product of ammonium assimilation causes a rapid and reversible inactivation of

nitrate transport (Flynn, 1991). Given this information, it is possible that when ammonium

concentrations reached zero, algal nitrate fixation took place. Although it has to be mentioned that

microalgae preference towards nitrogen sources is species dependent.

During the first 26 days (no nitrification), a calculation was performed to investigate the fate of the

incoming nitrogen. From day 12 to day 19 (7 days), the VSS concentration increased with 1.04 g VSS

L-1 and because the SRT was equal to the HRT, 5.12 g VSS was washed out with the effluent. 1.32 g

VSS d-1 was formed or in total, 9.27 g biomass was formed during this 7 day period. From the analysis

of filtered versus non-filtered reactor content, it was calculated that 10.4% (w/w) of biomass

comprised of Kjeldahl nitrogen. As a consequence, the formed biomass during these 7 days

contained in total 964 mg Kjeldahl nitrogen. This is 69% of all incoming nitrogen (50 mg N L-1 d-1 or

PART 4 Discussion

56

1400 mg N during 7 days). The remaining 31% was calculated to be in the effluent in the form of

Kjeldahl nitrogen (a reactor concentration of 180 mg N L-1 on average of which 50% accounted for

ammonium nitrogen and 50% organic nitrogen). To calculate the percentage of heterotrophic

biomass formed, COD removal was taken into account. During the 7 day period, on average 355 mg

COD d-1 was removed, partly in the reactor effluent, partly used for heterotroph biomass production.

47 mg COD d-1 was removed with the effluent, while 308 mg COD d-1 was removed by the

heterotrophs. Because 0.4 g dry weight is formed per g COD removed, 0.123 g VSS d-1 was formed.

Considered the part of this formed biomass that ends up in the effluent, 0.028 g VSS L-1 d-1 is

generated. During these 7 days, 0.78 g heterotroph biomass was formed which is 8.4% of total

biomass production. The production of this amount of biomass consumed only 39 mg N during this 7

days (0.05 g N assimilated for each g biomass produced).

There was no nitrification noticeable in the PBR, although a batch test (Results; paragraph 1.2)

showed a low biomass nitrification activity. The AOB in the reactor biomass had an activity of 7.33

mg NH4+-N g VSS-1 d-1 or a potential 12 mg NH4

+-N L-1 d-1 could be nitrified (with a reactor VSS

concentration of 1.5 g VSS L-1). In conclusion, there was a low but not sufficient potential for

nitrification of the incoming 50 mg NH4+-N L-1 d-1. Because dissolved oxygen in the reactor did not

reach levels higher than 1.4 mg O2 L-1 while the concentration in the batch test was kept higher than

5 mg O2 L-1, the shortage of oxygen was considered a possible cause of nitrification inactivity. In this

first period with periodically low oxygen concentrations, it was possible that the inner part of the

algal-bacterial flocs experienced anaerobic conditions and denitrification occurred. The mass balance

calculations revealed only small losses in the startup period. With a reactor oxygen concentration

lower than 2 mg O2 L-1, anaerobic cores are possible in flocs with a diameter of 400 µm or higher

(STOWA, 1999).

After 60 days of reactor operation, nitrification initiated. Nitrification efficiency increased and

reached a maximum value of 70% at day 125. This sudden initiation could be induced by adaptation

to the light conditions (see further).

Another calculation (for a 7 day period starting from day 103) was performed to investigate the fate

of the incoming influent nitrogen assuming a constant VSS concentration of 2 g VSS L-1. During this

period the nitrification efficiency was on average 52% and 8.4 g VSS was washed out with the

effluent. From the analysis of filtered versus non-filtered reactor content, it was calculated that 7.5%

(w/w) of biomass comprised of Kjeldahl nitrogen. Knowing this, the produced biomass contained 630

mg Kjeldahl nitrogen which is 30% of all incoming nitrogen (75 mg N L-1 d-1 or 2100 mg N during 7

days). The remaining 18% was calculated to be for 5% in the effluent in the form of Kjeldahl nitrogen

(a reactor concentration of 25 mg organic-N L-1 on average), 4% nitrate accumulation in the reactor

and the remaining 9% or 28 mg N d-1 was not accounted for. It is possible that this amount ended up

as nitrogen gas by denitrification.

In literature, only few recent studies report the use of an algal-bacterial consortium. Karya et al.

(2013) nitrified artificial wastewater (50 mg NH4+-N L-1) supported by photo-oxygenation in an algal–

bacterial consortium. Ammonium removal was primarily by nitrification (81–85%) rather than by

algal ammonium uptake. The nitrogen loading rate was 25 mg NH4+-N L-1 d-1 in comparison with the

PART 4 Discussion

57

71 – 91 mg NH4+-N L-1 d-1 we used. White light illumination was used instead of the blue-red enriched

light we used which can inhibit AOB and NOB activity (see further) and their cycles lasted for 12 and

24 hours instead of 8 hours. A settling time of 5 hours was used which retained almost all biomass

with low biomass growth rates as a consequence. In comparison with other studies concerning urine

nitrification and working with nitrifying sludge only (Literature review), only Chen (2009) achieved

100% nitrification of 10-29% diluted human urine, in a 26 L SBR configuration. Other studies

concerning urine nitrification for stabilization, reported 50% NO2- and 50% NH4

+ in the effluent with

the aim at partial nitritation/anammox or chemical nitrite oxidation experiments.

1.2. REACTOR OPERATION PARAMETERS

The initial nitrogen loading rate was increased from 50 mg N L-1 d-1 to 90 mg N L-1 d-1 during 125 days

with nitrification efficiencies going up to 69%. These loading rates are higher than the reported 25

mg NH4+-N L-1 d-1 of Karya et al. (2013), but nitrification efficiencies were lower. Although we dealt

with biomass washout, the reactor nitrification potential during its optimal nitrification period (50 mg

N L-1 d-1) was higher than the values obtained by Karya et al. (2013).

During each 8 hour cycle, dissolved oxygen concentration decreased after influent dosage until a

stable value was reached (variable during reactor operation, but always higher than 0.5 mg O2 L-1). At

that stable oxygen concentration, the algal oxygen production was in balance with the oxygen

consumption by nitrification. The oxygen concentration remained at this low value until all

ammonium and nitrite was oxidized after which it increased again to values between 6.55 and 8.88

mg O2 L-1 due to photosynthesis. Karya et al. (2013) observed a similar oxygen concentration profile

in their reactor, but due to a high oxygen production rate of 0.46 kg m-3 d-1, the oxygen concentration

increased until supersaturation (12 mg O2 L-1) after which it decreased to a stable 8 mg O2 L

-1. They

suggested that algal photosynthesis is carried out after ammonium depletion with nitrate as a

nitrogen source. This was confirmed by Wu et al. (2013).

During nitrification, the SVI was 58.06 ml g-1 which indicated a sufficiently well settling capacity.

Karya et al. (2013) reported an SVI of 160 ml g-1. Their higher SVI is explicable to the long settling

times used (up to 5 hours). In this way, badly settling microorganisms remained in the reactor

suspension. Van Den Hende et al. (2011) used MaB-flocs (Microalgal Bacterial flocs) with an SVI of 76

± 15 ml g-1, which increased rapidly when bicarbonate was dosed.

COD removal efficiencies fluctuated between 61 and 93% during the first period without nitrification

while values between 44% and 84% were obtained during the period with stable nitrification.

Maximum daily COD removal per gram VSS was 54 mg COD g VSS-1 L-1 d-1 (or 77.03 g m-3 d-1) in

comparison with 233 ± 57 mg COD gVSS-1 L-1 d-1 (Van Den Hende et al., 2011). In high rate algal

ponds (HRAPs), the average mean SBOD5 removal rate was 23.8 g m-3 d-1 with a HRT of 4 days as

reported by Park and Craggs (2010). Gutzeit et al. (2005) investigated the algal-bacterial regime in a

pilot scale SBR configuration and reported COD removal rates of 55.9 g COD m-3 d-1.

PART 4 Discussion

58

1.3. PHOTO-AERATION

From the dissolved oxygen profile (Results; paragraph 0), it was possible to estimate the nitrification

oxygen demand and algal oxygen production rate. From the moment ammonium and nitrite were

fully oxidized, the increasing oxygen concentration was the oxygen production rate (assuming that

other oxygen consuming processes were completed or negligible). The obtained oxygen production

rate was 0.134 mmol O2 g TSS-1 h-1 (or 1.36 mg O2 L-1 h-1) which is low compared to oxygen production

rates in fully optimized algae systems ranging from 4 – 6 mmol Lculture-1 h-1 (Drapcho and Brune, 2000;

Javanmardian and Palsson, 1992). Karya et al. (2013) reported an oxygen production rate of 19 mg L-1

h-1 which is considerable higher than 1.36 mg O2 L-1 h-1. The low oxygen production rate may be

caused by the low illuminated volume of less than 10% (> 90% of incidence light was absorbed in less

than 10 mm) in this non optimal photobioreactor. It has to be mentioned that the calculated oxygen

production rate can deviate from the effective value due to the headspace recirculation and

consequent re-aeration.

Nitrifying bacteria require, theoretically, 4.57 g of oxygen to convert 1 g NH4+-N to 0.95 g NO3

--N and

0.05 g biomass-N. Heterotrophic bacteria use oxygen to oxidize the organic carbon present to CO2. At

the moment of dissolved oxygen monitoring, 23 mg O2 L-1 was removed with COD and 20.7 mg O2 L

-1

was used to oxidize ammonium to nitrate during 1 cycle. The total theoretical oxygen consumption

during this cycle was calculated to be 113 mg O2 L-1. In contrary, the consumption and production

calculated based on the profile were 8.88 mg O2 L-1 and 10.9 mg O2 L-1 respectively. It has to be

mentioned that oxygen concentrations in the floc entities may reach 70 mg L-1 (Gutzeit et al., 2005)

which is not measurable with a normal DO-probe. As mentioned before, the true oxygen

consumption was possibly higher than noticeable because re-aeration occurred by headspace

recirculation. Photo-aeration provided enough oxygen for the nitrifiers which could be observed

from the oxygen profile (oxygen consumption was equal to oxygen production) and when at day 120

the compressed air was turned off the oxygen concentrations remained above 6.55 mg O2 L-1

(measured at the end of the cycle). Although the microalgal oxygen production was sufficient during

this period, there was not much potential for higher nitrogen loading rates. This was also observed

during reactor operation. When the nitrogen loading rate was increased to fast to 100 mg N L-1 d-1,

the reactor failed with a decreased activity of the AOB and NOB as a result.

Van Den Hende et al. (2011) calculated, based on a daily provided 154 mmol photons L-1 reactor and

assuming a realistic quantum yield of 0.08 (Emerson and Lewis, 1943), a maximum of 12.3 mmol O2 L-

1 reactor day-1 could be produced by photosynthesis. The measured SCOD removal rate was 10.0

mmol O2 L-1 reactor day-1.

To nitrify the target influent loading rate of 250 mg N L-1 d-1 and COD loading rate of 200 mg COD L-1

d-1 (33% urine), the algae need to obtain a specific oxygen production rate of 1.34 g O2 L-1 d-1 (or 56

mg O2 L-1 h-1). As mentioned earlier, algae can produce between 130 and 190 mg O2 L

-1 h-1 in optimal

designed photobioreactors which makes this target achievable. Figure 37 visualizes the ideal,

theoretical equilibrium between the microalgae and the nitrifiers starting from 100 units of nitrogen.

For every gram of ammonium nitrogen supplied to the algae, 5 grams of ammonium can be

PART 4 Discussion

59

converted to nitrate taking into account the necessary nitrogen for microalgae growth and oxygen

production.

Figure 37. Theoretical equilibrium between microalgae and nitrifiers starting from 100 units of

nitrogen and 90 units of COD (average ratio in urine).

1.4. REACTOR SHORT-COMINGS

1.4.1. ALGAE AND AOB/NOB GROWTH

Many parameters regulate algal growth and from all of them only light was limited by column

diameter. On the other hand, wavelengths were optimized for algal growth (blue and red light with

peak wavelengths of 460 and 630 nm respectively which are the absorption wavelengths of

chlorophyll pigments). A photon flux density of 300 µmol PAR m-2 s-1 was measured at the outer

reactor wall, but due to biomass TSS concentrations of 2.3 g L-1 light penetration depth was low. Light

attenuates exponentially as it penetrates into the culture medium estimated by Lambert-Beer’s law

(Lee, 1999). Figure 38 gives the light distribution inside a photobioreactor containing a biomass

concentration of 1 g L-1 with a light absorption coefficient of 200 m² s-1 and an illumination intensity

of 500 µmol PAR m-2 s-1 (Ogbonna and Tanaka, 2000). In our reactor the photon flux density is lower

and the biomass concentration higher, so a light penetration depth less than 1 cm can be assumed.

Nitrifiers

Microalgae

Heterotrophs

NH4

+-N NO3--N

Light

O2

O2

COD

90

100

80

20

80

84

365

CO2

640

427

PART 4 Discussion

60

Figure 38. Light distribution inside a photobioreactor containing 1 g L−1 Euglena gracilis cells with a

light absorption coefficient of 200 m2 kg−1. The photobioreactor was illuminated from one surface at

an intensity of 500 µmol m-2 s−1 (Ogbonna and Tanaka, 2000).

Other parameters, such as pH (6.5 – 7), nutrient quantity and quality (10% dilution of urine with 0.8 g

N L-1 and 0.05 g P L-1), stirring and temperature (20°C) were optimal for most microalgal species. Each

microalgal species has its own optimal parameter ranges, but by inoculating the reactor with a mix of

7 different microalgae species, some of them must prevail in this reactor environment. Under the

light microscope it was observed that from all inoculated species, Scenedesmus sp. predominated in

the end. Ji et al. (2013) reported optimal conditions for this genus that are similar to the reactor

conditions (a temperature range of 20 – 25°C and optimal pH between 5 and 10). Scenedesmus sp. is

also a fast growing species that can thrive in lots of environments (personal communication, van der

Steen). Next to Scenedesmus sp. (green algae), also Chlorella sp. (green algae), Synechocystis sp.

(cyanobacteria) and Leptolyngbya sp. (cyanobacteria) were observed. Because no pure cultures of

cyanobacteria were inoculated, it is likely that the two cyanobacterial species originated from the

undetermined algae we collected from a grassland pond.

Reported maximum specific growth rates at 20°C for AOB and NOB are 1.05 d−1 < μmax,AOB < 1.4 d−1

and 0.91 d−1 < μmax,NOB < 1.31 d−1 respectively (Munz et al., 2011). Specific algal growth rates in a

photobioreactor system are positioned between 0.53 and 0.84 d-1 (Miron et al., 2000). Although the

maximum specific growth rates are higher for the nitrifying bacteria, the algae were enriched in

comparison to the bacteria. This was probably due to photo-inhibition of the nitrifiers while the algae

were encountered with optimal conditions. The effective specific growth rate for the organisms is

influenced by parameters such as available nutrients and light and is related to the substrate affinity

constants (or half saturation constants; Ks). Table 19 gives an overview of different Ks-values for

nitrifying bacteria and microalgae found in literature. Although, values found in literature vary highly.

PART 4 Discussion

61

Table 19. Typical values of substrate affinity constants (Ks) and doubling time for AerAOB, NOB, and

Microalgae species (Guisasola et al., 2007; Hein et al., 1995; Lackner et al., 2008). ‘-‘: not applicable.

Parameter (Unit) AerAOB NOB Microalgae

Ks oxygen (mg O2 L-1) 0.6 2.2 -

Ks ammonium (mg N L-1) 2.4 - 0.01

Ks nitrite (mg N L-1) - 5.5 -

Ks nitrate (mg N L-1) - - 0.01

Ks inorganic C (mg C L-1) 21.4 - 13.2

Doubling time (d) 0.34 0.48 1

As ammonium concentrations were almost always higher than the half saturation constant of the

AOB and Ks;AOB > Ks;microalgae, the AOB were in the advantage because of their higher maximum growth

rates. Although, microalgal inorganic carbon affinity was higher while the CO2 concentrations were

low in the reactor, which favors the microalgae.

Another factor that encouraged algal growth was the washout of biomass. This was the consequence

of the inability to select for the largest flocs and initial short settling times. An important parameter

to this regard is the settling time which was initially chosen to be 10 minutes (with the last 8 minutes

effluent pumping). This short settling time was chosen because long settling times enhance the

retention of poorly settling sludge flocs. Despite this measure, no large flocs formed and thus no

sufficient settling occurred. Sludge retention time (SRT) was equal to the hydraulic retention time

(HRT) of 6.7 days and the effluent VSS concentration was equal to the reactor VSS concentration

which fluctuated between 1.42 and 3.03 g VSS L-1 during the first 55 days. Very good settling was

obtained after 15 days after inoculation with fresh ABIL (Ammonium Binding Inoculum Liquid;

Avecom). These good settling properties are linked to the high density of the new formed flocs, due

to the high calcium carbonate content of this ABIL sludge (Figure 39). This resulted in lower VSS/TSS

ratios.

Figure 39. Settling phase in the photobioreactor after reinoculation with ABIL and HANDS.

PART 4 Discussion

62

In conclusion, different factors influence the activity ratio between microalgae and nitrifying bacteria

of which competition towards ammonium and inorganic carbon is a very important one. It should be

possible to guide the activity ratio to a higher activity for the nitrifiers by means of light control

(sufficient volume of dark zone), and depending on half saturation constant of both groups, inorganic

carbon control. Supplying high enough carbon dioxide concentrations would favor the nitrifiers due

to their higher Ks value.

1.4.2. SETTLING

Initially no settling occurred. Only after reinoculation with ABIL, biomass settling was visible. After

133 days, the SVI5 was 90.32 ml g-1 and SVI30 was 58.06 ml g-1 which indicated a sufficiently well

settling capacity. This resulted in good settling biomass when the settling time was increased to 20

minutes. Next to the increased settling time, a leakage in the flow meter was detected and solved.

This leakage may have caused small amounts of air bubbling into the reactor hindering the settling.

To obtain good settling characteristics, the formation of aerobic granules or dense flocs is an option.

The microalgae could form the outer layer of the granule while the nitrifying bacteria would be

protected from light inhibition in the inner layers of the granule/floc. Nevertheless, granule

formation is a complex process which depends on many parameters including substrate composition,

organic loading, hydrodynamic shear force, feast-famine regime, feeding strategy, DO, reactor

configuration, SRT, cycle time and settling time (Liu and Tay, 2004). Many studies agree that

substrate loading rate and settling time are the most important factors influencing sludge

granulation (Chen and LaPara, 2008; Qin et al., 2004a; Wang et al., 2004). Strategies for settlement

improvement are discussed further.

Settling properties are also dependent on the present microorganisms. It was observed that with less

spherical shaped organisms, different biomass properties were obtained (personal communication;

van der Steen). Their reactor contained filamentous cyanobacteria (Anabaena sp.) which resulted in

dense flocs that settled very well with a clear supernatant. The SRT of the system was very high.

1.4.3. PHOTO-INHIBITION

Nitrification started after 55 days. This sudden sufficient nitrification activity could be the

consequence of adaptation of the nitrifying community to several parameters. One of these

parameters is light. It is general known that AOB and NOB are inhibited by light. For this reason all

the batch tests were performed by excluding light from the Erlenmeyer flasks with tinfoil. Abeliovich

and Vonshak (1993) investigated effects of light on Nitrosomonas europeana nitrification activity in

reservoir water and in defined medium. Exponentially growing cells were strongly light inhibited in a

defined medium while ammonia provided some protection against this effect. The cells were

exposed for 1 hour and totally inhibited nitrification for 4 days, with recovery beginning only on day

5. Alleman et al. (1987) researched the impact of light exposure on an enriched Nitrosomonas culture

based on the work of Hooper and Terry (1974) (Table 20). A first conclusion was that the range of

410 – 415 nm particularly appears to be responsible for Nitrosomonas inhibition. Skinner and Walker

(1961) determined coincidentally an absorption peak at 415 nm for a reddish cytochrome C-type

pigment typically found in Nitrosomonas. A second conclusion was that light induced inhibition was

PART 4 Discussion

63

uniquely obtained during a period without respiration (resting), when no nutrients were present. A

last conclusion concerns the full recovery of the bacteria which was observed after 6 – 10 h of dark

exposure with ammonium presence. This reversibility was blocked by chloramphenicol which

suggests that the recovery involves repair or replacement of a distressed enzyme(s). To block light

inhibition, light exposure should be avoided or these bacteria should be maintained in an active

respiration state. No records were found about a longer period of light irradiation and potential

adaptation to light inhibiting conditions.

Table 20. Nitrosomonas respiration evaluation (Alleman et al., 1987).

Presuming, the AOB in the biomass suspension contained Nitrosomonas sp. it is likely that they were

light inhibited in the reactor at the moment NH4+-N was depleted. No light dark-cycle was present

whereby the nitrifiers could potentially recover during the dark phase and the used light was

enriched with the inhibitory wavelengths (410 – 415 nm). Floc formation in which the nitrifier

community is protected from light inhibition, was not obtained. It is possible that after 55 days of

reactor operation, the bacteria adapted to these conditions of illumination and started to nitrify.

1.5. REACTOR IMPROVEMENT POSSIBILITIES

Because organic nitrogen made up 92% (on average) of the total nitrogen influent concentration,

urea hydrolysis is a factor to account for. Due to the time delay of urea hydrolysis, and the time

necessary to build up oxygen concentrations again (1.36 mg O2 L-1 h-1), a cycle of 8 hours did not

provide enough time to be ready for the next influent dosage with ammonium accumulation as a

consequence. Udert et al. (2003b) reported an ureolysis rate of 1820 g Nm-3 d-1 (standard deviation ±

110 g Nm-3 d-1) for untreated stored urine at 25°C. To hydrolyze the dosed 32.3 mg N L-1 cycle-1 at the

reported ureolysis rate of 75.8 mg L-1 h-1, 26 minutes is necessary at 25°C. This may be a little while

longer because the reactor operated at 20°C. This ureolysis time may be sufficient to cause problems.

Karya et al. (2013) worked with wastewater in which all nitrogen was present in the form of

ammonium, in this way a time delay was avoided. A possible solution to resolve the time delay of

ureolysis is the use of hydrolyzed urine. However, this may result in other reactor activity problems

related to the high salinity (60 – 70 mS cm-1 for fully hydrolyzed urine).

From a batch test at day 146, it was found that AOB consumed 67.35 mg NH4+-N gVSS-1 d-1 or the PBR

could nitrify 121.9 mg NH4+-N L-1 d-1 (VSS concentration in the reactor was 1.81 g VSS L-1). NOB

produced 47.85 mg NO3--N gVSS-1 d-1 or the PBR could produce 86.61 mg NO3

--N L-1 d-1. At a loading

rate of 90 mg N L-1 d-1 the potential was present to oxidize all ammonium in the influent. As the

PART 4 Discussion

64

dissolved oxygen concentration in the reactor was less than 1 ppm, oxygen production might be the

limiting factor. Another factor could have been the light inhibition towards AOB and NOB.

A possible solution to deal with a possible ureolysis time delay and in mean while deals with the

periodically low reactor oxygen concentrations and biomass washout, is the use of longer cycles with

longer settling times. A longer cycle provides more time to restore the oxygen concentration and to

nitrify the present ammonium while it prevents washout of nitrogen accumulating biomass and as a

consequence, more nitrogen will be available for nitrification. Attention has to be paid for potential

FA formation when the same nitrogen loading rate is applied with fewer cycles. The use of fewer

cycles a day offers the opportunity to apply a more explicit feast-famine strategy which is a favorable

condition for anaerobic granule formation.

As discussed earlier, light inhibiting conditions are present and AOB and NOB are likely hampered in

their nitrification activity. It may be possible to replace blue and red enriched light by blue filtered

light. In this way the inhibition wavelengths of the nitrifiers are avoided and algae may be restrained

in growth speed. Baba et al. (2012) tested the effect of monochromatic light on growth and

photosynthesis of Botryococcus braunii and concluded that based on photo energy supplied, red light

is the most efficient light source. Kim et al. (2013) investigated the effects of wavelength and

wavelength mixing ratios on Scenedesmus sp. growth and nitrogen removal. The microalgae

production rate was highest in white light (containing both 450–475 nm and 630–675 nm), followed

in order by red, blue, and green light. Therefore, white light is more appropriate for the growth of

microalgae than providing a single wavelength. It was observed that nitrogen removal rate was the

same for white light, red light, and blue light. Finally, mixes of red and blue light provided 50% higher

microalgae production rates than with white light or a single wavelength. From these findings it

seems interesting to prevent nitrifier inhibition and temper algae growth with monochromatic red

light. All together, the use of this kind of artificial operation conditions is not realistic in practice were

sunlight is the standard. A better option is to select for AOB and NOB species that can cope with light.

To protect the nitrification activity of AOB and NOB from photo-inhibition, a wide reactor diameter

was chosen. As a consequence, low microalgal specific oxygen production rates were observed. If

adaption towards light inhibition conditions appears to be possible for AOB and NOB, an

improvement in reactor design may be implemented with a narrower reactor column. In this way, a

larger reactor volume would be illuminated with more oxygen production as a result. Good settling

biomass becomes than a requirement because a higher photosynthesis yields higher biomass

concentrations and biomass washout has to be avoided.

The volumetric exchange ratio (VER) was 15% (0.6/4) during the entire reactor operation and a

hydraulic retention time (HRT) of 6.7 days was maintained. These conditions are unfavorable for

aerobic granulation. Granulation is benefited by a low HRT or high VER, causing poor-settling biomass

to wash out, and a high organic loading rate, ensuring sufficient new biomass growth (Wang et al.,

2004). It is possible to increase the VER to 50% by reducing the number of cycles from 3 to 2 and

increasing the influent flow rate to 2 L d-1. Consequently, the nitrogen loading rate at a 10% dilution

of urine would be immediately 250 mg N L-1 d-1 which was the photobioractor target value. When

PART 4 Discussion

65

increasing the VER, it may be interesting to look for less concentrated waste streams such as low

strength wastewater or digestates.

1.6. MICROSCOPY

The initial inoculum consisted of 7 different microalgae species, a microalgae mix acquired from a

grassland pond and activated HANDS sludge. The 7 microalgal species comprised Chlorella sp.,

Haematococcus sp., Desmodesmus sp., Ankistrodesmus sp., Pediastrum duplex, Chlorella vulgaris and

Nannochloropsis sp. Microscope observations showed that small flocs of micro-algae and bacteria

were formed. Based on these pictures it was possible to distinguish 4 species: Scenedesmus sp.

(green algae), Chlorella sp. (green algae), Synechocystis sp. (cyanobacteria) and Leptolyngbya sp.

(cyanobacteria). Between the large microalgae cells, the bacteria were present in low numbers. The

dominating specie was Scenedesmus sp., the other species present originated from the algae mix

acquired from the grassland pond.

Karya et al. (2013) inoculated with a pure culture of Scenedesmus quadricauda, but selection took

place towards cyanobacteria (undetermined). Van Den Hende et al. (2011) used microalgal flocs

(MaB-flocs) which contained Chlorella sp., Pediastrum sp., Phormidium sp. and Scenedesmus sp. This

illustrates that Scenedesmus sp. and Chlorella sp. are suitable organisms in a consortium to treat

different waste streams. In high rate algal ponds (HRAP’s) or raceways, where algal biomass is grown

as a by-product of wastewater treatment, various species are reported to be present. Mostly, mixed

algal cultures are used for wastewater treatment (e.g. Scenedesmus sp, Micractinium sp, Actinastrum

sp, Pediastrum sp, Coelostrum sp, Chlorella sp. and Ankistrodesmus sp). Park et al. (2011) states that

algal species control has yet to be achieved in wastewater treatment HRAPs and algal dominance and

species interactions are still poorly understood. Gutzeit et al. (2005) treated wastewater with algal-

bacterial flocs containing only Chlorella vulgaris.

2. MEMBRANE BIOREACTORS

Startup strategies for urine nitrifying membrane bioreactors were investigated. The first experiment

investigated the need for a constant electrical conductivity by means of one membrane bioreactor

initiated at a high salinity while another reactor was subjected to an adaptive strategy with low initial

salinity.

2.1. INFLUENCE OF THE MEDIUM

During storage and transport the urea in source-separated urine becomes hydrolyzed to ammonia

and bicarbonate by microorganisms with urease activity. This ureolysis is a fast process at a rate of

1820 g Nm-3 d-1 (standard deviation ± 110 g Nm-3 d-1) for untreated stored urine at 25°C (Udert et al.,

2003b). A consequence of ureolysis is a high carbonate concentration (16.26 g CO3 L-1) and as a result

an electrical conductivity up to 60 – 70 mS cm-1 for 100% hydrolyzed urine. This high salinity together

with a high nitrogen concentration limits the nitrification activity of unadapted AOB and NOB.

PART 4 Discussion

66

Because ureolysis is a necessary reaction before nitrification can occur and high salinities form a

challenge for the AOB and NOB, synthetic hydrolyzed urine was used during the experiment.

2.2. MBR 1: HIGH SALINITY START-UP STRATEGY

MBR 1 operated at an electrical conductivity of 70 mS cm-1 (conductivity of 100% hydrolyzed urine).

During the experiment, the flow rate was 0.5 L d-1 and the influent concentration was increased from

10% to 100% (decreasing dilution).

During startup, the biomass was inactive (ammonium accumulation) due to the salt shock. The initial

VSS/TSS ratio of 0.5 gVSS gTSS-1 was much lower then the initial value in MBR 2 of 0.75 gVSS gTSS-1

which means that a considerable part of VSS was submitted to imidiate lysis (due to the high salt

shock). When the AOB and NOB activity increased, a typical nitrogen species profile during start-up

of a nitrification system was observed (Elawwad et al., 2012; Jubany et al., 2008; Yu et al., 2011). This

profile is characterized by an initial ammonium accumulation followed by nitrite accumulation and

ending with the nitrite oxidation towards nitrate. The AOB reached nitrification activity faster than

NOB. This can be caused by the initial inhibiting effect of free ammonia (FA), or because NOB are

more sensitive to higher salinities. Liu et al. (2008) and Uemura et al. (2012) reported that NOB are

more sensitive to increased salinity then AOB.

During the first 10 days a lot of foam was formed in the reactor (Figure 40Figure 40. Foam formation

MBR 1.).

Figure 40. Foam formation MBR 1.

Many reasons are associated with foaming, including excess production of extracellular polymeric

substance (EPS) by activated sludge microorganisms under nutrient-limited condition (or other stress

conditions), generation of filamentous organisms and gas provided in aeration tanks or produced in

anoxic zones of aeration tanks (Pujol et al., 1991). Di Bella et al. (2011) discuss the role of EPS

concentration in MBR foaming and state that the operational conditions of an MBR plant promote

foam formation as a result of high MLSS concentration, high SRT and low food-micro-organism ratio

(F/M). Furthermore, the tank where the membrane is placed acts as a foam trapping unit. Adversely

to these described MBR’s, our system does not have high MLSS concentration and the food-

microorganism ratio is not low due to the concentrated urine influent. Stress conditions present a

better explanation for foaming. An important stress factor in our MBR was the high salinity. Reid et

PART 4 Discussion

67

al. (2006) indicate that high salinity greatly affects the physical and biochemical properties of

activated sludge, increasing soluble microbial product (SMP) and EPS concentrations. Stress induced

by the salt shock was the most plausible cause for the observed foaming. From the light microscope

images, it was possible to distinguish filamentous bacteria in small numbers. Nevertheless, these

images were made after 60 days when foaming almost disappeared. On that moment it was likely

that the community altered due to the different conditions it was used to.

Foam in MBR’s has also been observed in the absence of filamentous organisms, and in these

conditions, the quantity of foam has been reported as being related to the concentrations of the EPS

(Nakajima and Mishima, 2005). Filamentous and foam-forming microorganisms can increase the EPS

concentration in the mixed liquor and worsen the fouling and foaming phenomena (Meng et al.,

2006).

During operation, the electrical conductivity decreased from 70 mS cm-1 to an average value of 45 mS

cm-1. This was due to the reaction of bicarbonate (originated from ammonium bicarbonate) with

protons to form carbon dioxide and water.

On the moment ammonium and nitrite reactor concentrations were practically zero (day 35), influent

(10% dilution of urine) was dosed again at a nitrogen loading rate of 0.05 g N L-1 d-1. From that

moment onwards, it was possible to increase the loading rate from 0 to 0.19 g N L-1 d-1 in 19 days

while the electrical conductivity fluctuated between 42 and 51 mS cm-1. No further activity losses

were observed and full salinity adaptation was assumed.

2.3. MBR 2: INCREASING SALINITY START-UP STRATEGY

MBR 2 started at an electrical conductivity of 5 mS cm-1 (normal conductivity of the nitrifying sludge).

The influent flow rate was 0.5 L d-1 during the experiment and the influent concentration was

increased from 10% to 100% fully hydrolyzed urine (decreasing dilution).

The nitrogen loading rate was increased to 0.32 g N L-1 d-1 in 6 days. After that, nitrite and ammonium

accumulated. In that time the electrical conductivity increased to 15 mS cm-1 and foaming occurred

together with a decrease in VSS concentration. This event indicates that the nitrification activity of

the used nitrifying B-sludge was not influenced by the increasing salt concentrations until 15 mS cm-1

was reached which appears to be the maximum electrical conductivity at which non-adapted

nitrifiers show nitrification activity. After the adaptation period of AOB and NOB to the high salinity,

which lasted for 16 days, the nitrogen loading rate was increased further until 0.500 g N L-1 d-1 was

reached. This loading rate was attained with an average urine influent nitrogen concentration of 8 g

N L-1 (100% hydrolyzed urine) and a flow rate of 0.5 L d-1. Electrical conductivity stagnated on 45 mS

cm-1. When the electrical conductivity reached 15 mS cm-1, foaming started. This proves the before

mentioned statement that this foam formation is stress related.

Batch tests were performed during the adaptation period and during the period that MBR 2 was

operated with 100% hydrolyzed urine. The nitrification activity acquired from these test during

adaptation correlated to the nitrification activity inside the reactor. For the second activity test,

which was performed 1 day before the reactor experienced 1 time the HRT of 16 days, the AOB and

PART 4 Discussion

68

NOB activities were respectively 36 mg NH4+-N L-1 d-1 and 106 mg NO3

--N L-1 d-1. The AOB activity did

not correlate with the reactor activity which was operated at a nitrogen loading rate of 500 mg N L-1

d-1. The day after the activity test however, ammonium accumulated which indicated the insufficient

AOB nitrification activity. The MBR experienced only for a short moment activity loss after which it

proceeded with the normal influent of 100% hydrolyzed urine.

2.4. COMPARISON WITH LITERATURE

Chen (2009) achieved full nitrification of urine using an SBR configuration featuring aerobic granules,

which treated 29% diluted human urine (1802 mg COD L-1 and 1515 mg NH4+-N L-1). An average

removal efficiency of COD and NH4+-N of 90, and 99.7% was obtained at the COD and TKN loading

rates of 1.2 g L-1 d-1 and 1.05 g TKN L-1 d-1, respectively. After 1.5 months of urine storage, the urine

was used into the SBR with a NH4+-N/TKN ratio in the influent as low as 0.42, resulting in a large

amount of TKN remained in the effluent due to insufficient activity of heterotrophs. The start-up

phase was conducted with 10% diluted human urine and the final nitrogen loading rate was reached

after 65 days. No record was found about the reactor salinity.

Sun et al. (2012) researched the possibility to stabilize human urine biologically using two SBR’s

which were inoculated with nitrifying bacteria and aerobic granules, respectively. The used urine was

stored for 10 – 15 days at 20°C after which the pH in the urine was above 9 and NH4+-N content was

more than 5.8 g N L-1. Reactor ammonia volumetric loading rate was 0.5 g N L-1 d-1 but no stable

nitrification was obtained. Nitritation reached a performance of 90% with nitrite accumulation and

NOB inhibition as a consequence.

Udert et al. (2003a) discussed the possibilities of a CSTR and SBR configuration in the nitrification of

source-separated urine. In the CSTR (continuous flow stirred tank reactor), nitrite ammonium

solutions were produced with an SRT of 4.8 days at 30°C. In the SBR, the SRT was more than 30 days.

Nitrate build-up was negligible in both reactors and nitritation rates were 0.780 g N L-1 d-1 in the CSTR

and 0.280 g N L-1 d-1 in the SBR. For the CSTR, the ammonia loading rate was increased from 0.450 g

N L-1 d-1 to 1.580 g N L-1 d-1. They conclude that the lack of nitrite oxidizers was largely due to nitrous

acid inhibition, hydroxylamine and high salt concentrations.

No publications have been found that report nitrification reactors operating at an electrical

conductivity of 45 mS cm-1 or higher.

2.5. ADAPTATION TO HIGHER SALINITIES

The inactive period, during adaptation to higher salinities, lasted for MBR 1 (at 60 – 70 mS cm-1) 12

days for the AOB and 23 days for the NOB while it took 16 days for the biomass in MBR 2 (at

electrical conductivities of 15 – 27 mS cm-1). Biomass of MBR 2 never completely lost its nitrification

activity. After adaptation, MBR 1 needed 22 days to reach an influent concentration of 45% while

MBR 2 needed 21 days to reach an influent concentration of 100%. This proves that a gradual

increasing nitrogen loading rate (and electrical conductivity) is beneficial for a fast start-up.

PART 4 Discussion

69

From the last nitrification activity test performed at different electrical conductivities it was seen that

the AOB from MBR 1 and 2 showed highest activity at an electrical conductivity of 15 mS cm-1. This

means that the optimal electrical conductivity shifted towards higher values (initial optimal and

normal conductivity was 5 mS cm-1) because 15 mS cm-1 was the initial inhibitory salinity. The NOB in

contrary behaved differently. NOB from MBR 1 were most active at low electrical conductivities (5

mS cm-1) while NOB from MBR 2 experienced optimal activity at the highest conductivity (45 mS cm-1)

with a small difference towards the lower salinities. This may be explained by the different start-up

strategies whereby NOB from MBR 2 adapted more gradually to higher salinities.

2.6. MICROSCOPY

The sludge flocs from MBR 1 and MBR 2 have an irregular form and a compact structure which

indicates the good settleability. The compact structure can be caused by the bubble aeration

(STOWA, 1999). The flocs coming from MBR 1 contain few filamentous bacteria while the sludge flocs

originating from MBR 2 are formed around filamentous bacteria. The flocs from MBR 2 contain a

large number of cauliflower-like structures. These formations have the typical morphology of

ammonium oxidizing bacteria clusters.

Bacteria form flocs in nutrient limited conditions, to protect themselves from washout. In aerated

systems, biomass (M) will increase and nutrients (F) decrease in a logarithmic manner. The F/M ratio

decreases rapidly and bacterial flocs form. The flocs become more solid when the sludge loading rate

(F/M ratio) decreases. The flocs from MBR 2 are more dense then the flocs from MBR 1. This can be

explained by the lower loading rate in MBR 1 in comparison to MBR 2.

3. INFLUENCE OF SALT CONCENTRATION ON NITRIFICATION ACTIVITY

To investigate the effect of salinity on nitrifying activity, nitrifying HANDS sludge was exposed to

different salinities (different concentrations of sodium chloride) (Table 21). During the preparation of

the biomass, the supernatant after centrifuging had a conductivity of 14.66 mS cm-1. Consequently,

highest nitrification activity was expected at the highest salt concentration of 5 g L-1 NaCl (13.60 mS

cm-1) because this was the normal working salinity of the HANDS sludge. This expectations were

confirmed for the NOB which had a nitrification activity of 8.57 mg NO3-- N g VSS-1 d-1 at a salt

concentration of 1 g L-1 NaCl and 13.88 mg NO3-- N g VSS-1 d-1 at a salt concentrations of 5 g L-1 NaCl.

The AOB did not show nitrification activity at any salt concentration. During the experiment, free

ammonia and free nitrous acid (HNO2) were at highest 0.10 and 0.026 mg N L-1 (Anthonisen et al.,

1976), which is too low to inhibit the nitritation. The AOB probably lost their nitrification activity

during storage at 4°C or the acquired HANDS was too old.

High percentages of salt are known to disturb the correct operation of conventional aerobic

wastewater treatment processes above chloride concentrations of 5 – 8 g L-1 (Ludzack and Noran,

1965). Several studies have been recently carried out on treating saline wastewater. Di Bella et al.

(2013) reported ammonium removal efficiencies to drop from 99% to 80% inside a membrane

PART 4 Discussion

70

bioreactor, as consequence of a salt concentration of 10 g NaCl L-1 (or 16.95 mS cm-1). Johir et al.

(2013) studied the gradual increase of salt concentration (0 to 35 g NaCl L-1 in 110 days) on the

performance of a membrane bioreactor. An increase of salt concentration from 0.5 to 10 g NaCl L-1

showed decreasing removal efficiencies of ammonia from 90 to 76%. When salt concentration

reached 35 g NaCl L-1, the specific ammonium removal efficiency decreased from an initial value of

8.2 to 0 mg NH4+-N g VSS-1 d-1 (Figure 41).

Figure 41. Profile of specific removal of organics (DOC) and NH4+-N at different salt concentrations

(Johir et al., 2013).

Lefebvre and Moletta (2006) stated by means of a literature review that biological treatment of

wastewater has proven to be feasible at high salt concentrations but that its efficiency depends on

adaptation of the biomass or the use of halophilic organisms. In the discussed studies, the activity of

the microorganisms always decreases starting from a salt concentration of around 10 – 15 g NaCl L-1.

A similar result was obtained in MBR 2.

The nitrification activity in MBR 1 and MBR 2 were followed up at different moments and at different

salt concentrations (Table 21). A first activity test resulted in very low activities for MBR 1 (4.32 mg

NH4+-N gVSS-1 d-1 and 1.42 mg NO3

--N gVSS-1 d-1 at the normal reactor conductivity of 70 mS cm-1).

The activities of MBR 2 were higher (11.45 mg NH4+-N gVSS-1 d-1 and 24.35 mg NO3

--N gVSS-1 d-1 at the

present reactor conductivity of 27 mS cm-1). These results agree with the existing conditions; MBR 1

was still recovering from the salinity shock while MBR 2 was adapting to a load of 250 mg N L-1 d-1.

For MBR 2, the AOB did not show a significant difference in nitritation activity at the conventional

salt concentration of 5 mS cm-1 which implies that AOB do not possess fast adaptation potential

because the low salinity did not restore nitrification activity. Another possibility is that the optimal

salinity for AOB activity shifted towards higher salinities. In that case the activity should be lower at

lower salinity which was not the case. The NOB on the contrary had a nitratation activity that was

double as high at 5 mS cm-1 which implies that NOB can adapt to higher salinity without a salinity

optimum shift. The NOB recovered approximately to their original nitrification activity. Liu et al.

(2008) and Uemura et al. (2012) reported that NOB are more sensitive to increased salinity then

AOB. No studies were found that imply a faster adaptation potential of NOB in comparison to AOB.

A second activity test (after 55 days, when MBR 1 was adapting to a load of 35% hydrolyzed urine

and MBR 2 was running on 100 % hydrolyzed urine) resulted in higher nitrification activities at all

PART 4 Discussion

71

salinities due to the advanced halotolerance. The NOB activities from MBR 1 confirmed the

observation of previous batch test that NOB can adapt to higher salinity without a salinity optimum

shift. However, the NOB of MBR 2 did not follow this trend but highest activity was observed at 45

mS cm-1. It is possible that it takes the NOB longer than the AOB to shift their optimum to higher salt

concentrations. Concerning the AOB activity in both reactors, AOB were more active in MBR 1

compared to their activity in MBR 2, although MBR 1 had a nitrogen loading rate of 0.180 g N L-1 d-1 in

comparison with the loading rate of 0.500 g N L-1 d-1 in MBR 2. This discrepancy in activities

originated from activity loss which was confirmed the day after this batch test. After the reactor

operated for 16 days (1x HRT) at 0.500 g N L-1 d-1, an ammonium concentration in the reactor of 300

mg NH4+-N L-1 was observed (after 57 days). This short activity loss could be due to a temporarily

lower dissolved oxygen concentration. It was observed that AOB from MBR 1 and 2 showed highest

activity at an electrical conductivity of 15 mS cm-1. This means that the optimal electrical conductivity

shifted towards higher values (initial optimal and normal conductivity was 5 mS cm-1 and 15 mS cm-1

was the initial inhibitory salinity).

Table 21. Nitrification activities at different salinities.

Biomass pH EC

(mS cm-1

)

AOB activity

(mg NH4+-N gVSS

-1 d

-1)

NOB activity

(mg NO3-- N gVSS

-1 d

-1)

HANDS 6.68 ± 0.09 4.54

0 8.57 ± 0.10

HANDS 6.67 ± 0.13 6.55 0 9.84 ± 1.38

HANDS 6.65 ± 0.13 9.95 0 9.91 ± 1.28

HANDS 6.63 ± 0.15 13.60 0 13.88 ± 0.26

HANDS + PBR supernatant 7.34 ± 0.45 9.45 27.9 12.0

B-sludge initial 7.55 ± 0.29 4.54 ± 0.01 21.3 ± 0.91 59.9 ± 3.49

MBR 1 (day 21) 7.09 ± 0.07 11 ± 0.57 1.95 0.60

MBR 1 (day 21) 6.85 ± 0.12 66 ± 2.55 4.32 1.42

MBR 1 (day 55) 7.00 ± 0.03 5.82 ± 0.03 64.88 ± 6.85 88.80 ± 5.77

MBR 1 (day 55) 7.00 ± 0.03 15.08 ± 0.49 77.93 ± 7.88 87.96 ± 6.64

MBR 1 (day 55) 6.99 ± 0.03 45.20 ± 0.14 67.03 ± 6.55 43.97 ± 5.53

MBR 2 (day 21) 7.09 ± 0.08 9.66 ± 0.54 11.12

49.13

MBR 2 (day 21) 6.94 ± 0.07 24.5 ± 0.28 11.45

24.35

MBR 2 (day 55) 7.07 ± 0.02 6.18 ± 0.02 36.24 ± 1.02 104 ± 5.28

MBR 2 (day 55) 6.98 ± 0.03 16.04 ± 1.01 46.39 ± 3.78 97.98 ± 8.07

MBR 2 (day 55) 6.94 ± 0.04 44.67 ± 1.14 37.95 ± 2.96 113.31 ± 5.96

The possible capability for quick adaptation of the NOB can be seen from an activity test after 44

days at elevated temperatures. The activity of biomass from MBR 2 was tested at 35 mS cm-1 at a

temperature of 34°C and at 40°C while the biomass was used to nitrify at 20°C. The results are

PART 4 Discussion

72

presented in Table 22. The AOB activities did not change much in comparison with the activity 11

days later (at day 55; Table 21), but the NOB activity increased severely.

Table 22. Nitrification activity MBR 2 at different temperatures (after 44 days).

Temperature AOB activity (mg

NH4+-N gVSS

-1 d

-1)

NOB activity (mg

NO3--N gVSS

-1 d

-1)

MBR 2 (34°C) 38 ± 1 195 ± 9

MBR 2 (40°C) 28 ± 5 179 ± 8

4. INFLUENCE OF SALINITY AND AMMONIUM CONCENTRATION ON THE

MICROALGAL GROWTH

Besides the effect of salt concentration also the effect of different ammonium concentrations on the

microalgal growth was analyzed. The results showed that each algal species has different optimal salt

and ammonium concentrations.

Ankistrodesmus sp. favored low salt concentrations while Nannochloropsis sp. and Chlorella vulgaris

preferred higher salt concentrations. No real trends were noticeable for the other species. Growth at

high salt concentrations for Ankistrodesmus sp. was inhibited, as was visible through the lag phase.

These results were confirmed by Mohapatra et al. (1998) who studied the effect of salinity on the

growth of the fresh water algae Scenedesmus sp. They concluded that in normal pond water the

optimal growth was reached at a salt concentration of 0.42 g NaCl L-1 and further salinity increase

caused growth retardation. In nutrient enriched pond water however, the maximum growth

occurred at 1.69 g NaCl L-1. Kaewkannetra et al. (2012) also studied the effect of salinity (2.9 – 175 g

NaCl L-1) on the cultivation of Scenedesmus obliquus (for biodiesel production) and highest growth

rate was observed at the lowest salinity.

The mix of algae species showed better growth than the individual species. Only the mix of algae

reached the exponential growth phase. Optical densities at 620 nm between 0.65 and 1.25 were

reached depending on ammonium and salt concentration and no optical densities higher than 0.4

were observed for the single species.

Concerning the ammonium concentration, again highest growth was noticed for the algae mix (OD at

620 nm ranging from 0.60 – 1.10 depending on ammonium concentration). The lowest concentration

(50 mg NH4+-N L-1) had a positive effect on the growth of the algae mix in comparison with the higher

concentrations (final OD of 1.10 in comparison of a final OD between 0.67 and 0.80 respectively).

With reference to the single species, Chlorella sp., Haematococcus sp. and Ankistrodesmus sp.

preferred low ammonium concentrations while Nannochloropsis sp. preferred high ammonium

concentrations. This may be due to the effect of the higher salinity at higher ammonium

concentrations since Nannochloropsis sp. is a marine microalgae and favors higher salinities.

PART 4 Discussion

73

In conclusion, these microtiter growth tests were a rough means to gather information about salt

and ammonium concentration effects on algae on a short time notice and in high throughput. It was

clear that a mixed community of microalgae showed better growth in all conditions than the pure

cultures.

5. ECONOMICAL POINT OF VIEW

The primary energy consumption for the conventional nitrification/denitrification process is 45 MJ

kg–1 N (and 109 MJ kg–1 N for Nitrification/denitrification with methanol as substrate), compared with

37 - 45 MJ kg–1 active nitrogen produced with the Haber-Bosh process (Maurer et al., 2003).

Depending on an analysis of both, recovery process versus removal and recapture process (or

‘recycling over the atmosphere’), it has to be determined which approach is economically most

feasible for urine treatment.

Maurer et al. (2003) gives an overview of the possible technologies for nutrient recovery from liquid

wastes and compares them with the current practice. In this story, we fit in biological nitrification to

stabilize source separated urine and facilitate storage and transport with avoidance of nitrogen

losses. The conventional nitrification/denitrification and the subsequent Haber-Bosch process

together have typical operating energy demand of 90 MJ kgN-1. However, elimination of nitrogen in a

Sharon-Anammox process (19 MJ kgN-1) in combination with fertilizer production (45 MJ kgN-1) is

more favorable (64 MJ kgN-1). Source separation of urine, stabilization and nutrient recovery in the

form of struvite has an energy demand of 102 MJ kg–1. In struvite, every kilogram nitrogen comes

with 2.2 kilogram phosphorus and to produce an equivalent amount of P-fertilezer, 64 MJ is required.

When considering the possible advantage of nutrient recovery, it is important to take into account

infrastructural costs for source separating toilets, piping and storage vessels together with the

transportation costs.

When comparing high rate algal ponds (HRAP) with conventional wastewater treatment, high rate

algal ponds are cheap in operation and construction, but a high land area is required and algae

harvesting costs can be high. The HRT in HRAP’s is situated between 4 and 10 days. This means that

the entire volume of wastewater from 4 – 10 days needs to be stored in a pond with a depth of on

average 0.5 m. When the area is available, the algae settle well and the biomass can be used as a

high quality fertilizer or for biofuel production, HRAP’s can be feasible. No detailed economical study

was carried out.

6. CONCLUSIONS

Biological oxidation of all nitrogen present in urine to nitrate is a promising pretreatment step for

further nutrient recovery processes. In this way the urine can be stabilized for further processing, no

nitrogen can escape by volatilization of ammonia and odor problems are avoided.

PART 4 Discussion

74

Photo-aeration offers a possibility to reduce the aeration costs for the biological nitrification of urine.

A consortium of microalgae and nitrifying bacteria was developed in which the microalgae produced

enough oxygen in-situ to oxidize ammonium to nitrate. Further research is necessary to optimize the

operational conditions for the photobioreactor.

An optimal startup strategy for the urine nitrification process in a membrane bioreactor

configuration was implemented with success. An important element hereby was the adaptation of

AOB and NOB to high salinities linked to fully hydrolyzed urine, which was achieved. Gradual

adaptation appeared to be the best strategy to startup such a system.

PART 4 Discussion

75

7. FURTHER RESEARCH

Some suggestions for possible further research are mentioned here.

7.1. PHOTO-AERATION FOR URINE STABILIZATION

A high possibility exist that the photobioreactor operation was hampered during the first 55 days due

to light inhibition towards the nitrifiers. In this regard, it would be interesting to research the

influence of light sensitivity and adaptation towards illumination on the operation of a nitrifying

photobioreactor. A distinction between AOB and NOB and long term and short term irradiation could

be made.

Further research should be conducted on the equilibrium relationship between microalgae and

nitrifying bacteria. Parameters that influence the ratio of microalgae towards nitrifying bacteria and

AOB and NOB mutual should be further identified and control strategies should be tested. When a

full understanding of the consortium and the ratio determining mechanisms is established, it is

possible to operate the photobioreactor in a desirable manner.

Although good biomass settling characteristics were obtained, further research should determine the

possibility to form denser flocs or even microalgal-bacteria granules. This is potentially important to

offer light protection for the nitrifiers and to further improve settling.

7.2. URINE STABILIZATION WITH AN MBR CONFIGURATION

In the timeframe of this thesis, one startup strategy for urine nitrifying MBR’s was tested. Further

research should be conducted towards other possible startup strategies. Different possibilities are

shortly listed:

The comparison of a startup with an increase in influent flow rate versus an increase in

concentration (= decreasing dilution) to evaluate the fastest way of achieving a target

volumetric conversion rate.

The same test could be done with fresh human urine instead of with synthetic composed

urine. This experiment should be conducted with and without preceding hydrolysis to

investigate the effect of ‘slow down’ of urea hydrolysis.

It is possible to investigate the behavior of different kinds of inoculum sludge towards the

high salt concentrations. The effect of inoculation with activated sludge from a saline

environment (obtained from a marine environment, from aquaria or from the industry) could

be investigated.

Next to these different startup strategies, it is interesting to molecularly evaluate the membrane

bioreactor biomass towards shifts in bacterial community during operation under high salt

concentrations. In this way it is possible to identify the AOB and NOB species that are best suited for

treatment of saline and/or concentrated waste streams.

References

76

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Appendix

86

APPENDIX

1. EFFECT OF AMMONIUM CONCENTRATION ON DIFFERENT MICROALGAL SPECIES

0

0,05

0,1

0,15

0,2

0,25

0,3

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L0

0,05

0,1

0,15

0,2

0,25

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L

B. Haematococcus sp.

0

0,05

0,1

0,15

0,2

0,25

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L

C. Desmodesmus sp.

0

0,1

0,2

0,3

0,4

0,5

0,6

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L

D. Ankistrodesmus sp.

0

0,02

0,04

0,06

0,08

0,1

0,12

0,14

0,16

0,18

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L

E. Pediastrum duplex

0

0,05

0,1

0,15

0,2

0,25

0,3

0,35

0,4

0,45

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L

F. Chlorella vulgaris

A. Chlorella sp.

Appendix

87

Figure 42. Growth curves of different microalgae species at different ammonium concentrations. A.

Chlorella sp.; B. Haematococcus sp.; C. Desmodesmus sp.; D. Ankistrodesmus sp.; E. Pediastrum duplex;

F. Chlorella vulgaris; G. Algae mix from grassland pond; H. Nannochloropsis sp.

2. EFFECT OF SALT CONCENTRATION ON DIFFERENT MICROALGAL SPECIES

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0,8

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

G. Algae mix from pond

0

0,02

0,04

0,06

0,08

0,1

0,12

0,14

0 50 100 150

OD

at

62

0 n

m

Time (h)

50 mg NH4-N/L

100 mg NH4-N/L

250 mg NH4-N/L

1000 mg NH4-N/L

H. Nannochloropsis sp.

0

0,05

0,1

0,15

0,2

0,25

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

A. Chlorella sp.

0

0,02

0,04

0,06

0,08

0,1

0,12

0,14

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

B. Haematococcus sp.

0

0,04

0,08

0,12

0,16

0,2

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g NaCl/L

2 g NaCl/L

3.5 g NaCl/L

5 g NaCl/L

C. Desmodesmus sp.

0

0,05

0,1

0,15

0,2

0,25

0,3

0,35

0,4

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

D. Ankistrodesmus sp.

Appendix

88

Figure 43. Growth curves of different microalgae species at different salt concentrations. A. Chlorella

sp.; B. Haematococcus sp.; C. Desmodesmus sp.; D. Ankistrodesmus sp.; E. Pediastrum duplex; F.

Chlorella vulgaris; G. Algae mix from grassland pond; H. Nannochloropsis sp.

0

0,02

0,04

0,06

0,08

0,1

0,12

0,14

0,16

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

E. Pediastrum duplex

0

0,05

0,1

0,15

0,2

0,25

0,3

0,35

0,4

0,45

0,5

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

F. Chlorella vulgaris

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0,8

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

G. Algae mix pond 0

0,02

0,04

0,06

0,08

0,1

0,12

0,14

0,16

0,18

0 50 100 150

OD

at

62

0 n

m

Time (h)

1 g/L NaCl

2 g/L NaCl

3.5 g/L NaCl

5 g/L NaCl

H. Nannochloropsis sp.

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